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w a t e r r e s e a r c h 4 6 ( 2 0 1 2 ) 2 5 9 e2 6 6
Available online at www.sciencedirect.com
journal homepage: www.elsevier.com/locate/watres
A review of water recovery by vapour permeation through membranes Brian Bolto*, Manh Hoang, Zongli Xie CSIRO Materials Science and Engineering, Private Bag 33, Clayton South MDC, Victoria 3169, Australia
article info
abstract
Article history:
In vapour permeation the feed is a vapour, not a liquid as in pervaporation. The process
Received 26 May 2011
employs a polymeric membrane as a semi-permeable barrier between the feed side under
Received in revised form
high pressure and the permeate side under low pressure. Separation is achieved by the
21 October 2011
different degrees to which components are dissolved in and diffuse through the
Accepted 24 October 2011
membrane, the system working according to a solution-diffusion mechanism. The mate-
Available online 31 October 2011
rials used in the membrane depend upon the types of compounds being separated, so water transport is favoured by hydrophilic material, whether organic or inorganic. The
Keywords:
process is used for the dehydration of natural gas and various organic solvents, notably
Vapour permeation
alcohol as biofuel, as well as the removal of water from air and its recovery from waste
Pervaporation
steam. Waste steam can be found in almost every plant/factory where steam is used. It is
Gas drying
frequently contaminated and cannot be reused. Discharging the spent steam to the
Solvent dehydration
atmosphere is a serious energy loss and environmental issue. Recycling the steam can
Steam recovery
significantly improve the overall energy efficiency of an industry, which is responsible for massive CO2 emissions. Steam separation at high fluxes and temperatures has been accomplished with a composite poly(vinyl alcohol) membrane containing silica nanoparticles, and also, less efficiently, with an inorganic zeolite membrane. Crown Copyright ª 2011 Published by Elsevier Ltd. All rights reserved.
Contents 1. 2.
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Water permeability of membrane systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.1. Dehydration of natural gas . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2. Drying of compressed air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3. Flue gas dehydration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4. Dehydration of ethanol . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.5. Dehydration of isopropanol . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.6. Dehydration of acetonitrile . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.7. Steam recovery . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.8. Miscellaneous water removal applications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
* Corresponding author. Tel.: þ61 3 9545 2037; fax: þ61 3 9545 1128. E-mail address:
[email protected] (B. Bolto). 0043-1354/$ e see front matter Crown Copyright ª 2011 Published by Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2011.10.052
260 260 261 261 261 261 262 262 262 263
260
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3. 4. 5.
Membranes used in vapour permeation processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Comparison of vapour permeation and pervaporation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1.
Introduction J¼
Vapour permeation employs a polymeric membrane as a semi-permeable barrier between a feed side under high pressure and a permeate side under low pressure (Brinkmann et al., 2003). Unlike pervaporation, the feed is a vapour and not a liquid, so there is no phase change or significant temperature difference across the membrane. Separation is achieved by the different degrees to which components are dissolved in and diffuse through the polymer. The polymers used in the membrane will hence depend on the types of compounds being separated (Leemann et al., 1996). The driving force can be approximated to the difference in partial pressures of the components in the feed. Compared to pervaporation, vapour permeation effectively increases the permselectivity of water (Fan et al., 2002) and is capable of higher fluxes (Sander and Janssen, 1991). Another advantage is that the membrane area required is less, and there is a lower likelihood of membrane damage by impurities in the feed. Although vapour permeation operates with the same type of membranes as pervaporation, it has taken longer for its technical realisation. The first industrial vapour permeation plant was built in Germany in 1989 for dehydrating 30 kL/d of ethanol (Sander and Janssen, 1991). More than 100 membrane vapour-gas separation systems have been installed worldwide for recovering high value solvents, liquefied petroleum gas, refrigerant gases, monomers such as ethylene, propylene and vinyl chloride, and for removing the acid gases H2S and CO2 from natural gas and hydrocarbon vapours from air streams in the petrochemical industry (Baker et al., 1998; Jonquie`res et al., 2002). The transport of water vapour and inert gases through polymeric membranes was reviewed in the literature some time ago (Metz et al., 2005). The authors noted that the use of membranes for this purpose is of major industrial importance, with applications in areas such as the drying of natural gas and compressed air, protective apparel, packaging materials, roofing covers and humidity control in confined spaces, involving air conditioning in buildings, aviation and space flight. Steam recovery is also in this category. Vapour permeation, like pervaporation, works according to a solution-diffusion mechanism. The molecular interaction between the membrane and the separated species is the determining factor rather than the molecular size. The main component of the process is the membrane material which determines the permeation and selectivity and hence the separation properties of the process. The driving force for the mass transfer of permeate from the feed side to the permeate side of the membrane is the chemical potential gradient established by applying a difference in the partial pressures of permeate across the membrane (Aptel et al., 1972; Neel, 1991). The membrane performance is expressed as the membrane flux J in
263 263 265 265
P$DP l
where P is the permeability coefficient for a vapour, l is thickness of the membrane, and DP is the difference in pressure between the feed and permeate sides (Scott, 1998).
2.
Water permeability of membrane systems
The water vapour and nitrogen permeability behaviour of 19 polymers reported in the literature has been summarised (Metz et al., 2005). The data extrapolated to zero water activity are given in Table 1. Many of the values were obtained from pure gas permeabilities by calculating the ratios for the mixtures. The permeabilties are given in Barrer, a non-SI unit of gas permeability (specifically, oxygen permeability) used in the contact lens industry (Alter, 1962): one Barrer ¼ 1011 (cm3 O2) cm cm2 s1 mmHg1. In real mixtures however, water may swell the membrane so that its effect on the slower
Table 1 e Water vapour permeabilities and water/ nitrogen selectivities at 30 C for various organic polymers (Metz et al., 2005; Sijbesma et al., 2008). Polymer
PEBAX 1074a PBT/PEO block copolymerb Sulphonated poly (ether ether ketone) Polydimethylsiloxane Sulphonated poly (ether sulphone) Ethyl cellulose Cellulose acetate Poly(phenylene oxide) Poly(ether sulphone) Natural rubber Polysulphone Polycarbonate Polystyrene Polyimide Polyacrylonitrile Poly(vinyl chloride) Polyamide 6 Polypropylene Poly(vinyl alcohol) Polyethylene
Water permeability, barrer
Selectivity, H2O/N2
160,000 85,500 61,000
200,000 40,500 10,200,000
40,000 15,000
143 214,000
20,000 6000 4060 2620 2600 2000 1400 970 640 300 275 275 68 19 12
6060 24,000 1068 10,500 299 8000 4670 388 5,330,000 1,880,000 12,500 11,000 227 33,300 6
a PEBAX 1074 is a blend of polyether block amide (nylon 12) and poly(ethylene oxide). b (Polybutylene terephthalate)/poly(ethylene oxide) block copolymer.
261
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species is not known with certainty. Generally, for binary mixtures of permanent gases a higher selectivity is accompanied by a lower permeability, but such a relationship does not hold for water vapour in a mixture with a permanent gas, where most of the highly selective polymers also have a very high permeability. There can be a variation of over seven orders of magnitude in selectivity and five orders of magnitude in permeability. The effect of polymer structure on performance is variable. Completely non-polar membranes such as those based on natural rubber, polystyrene, polypropylene and polyethylene tend to have poorer permeabilities. The low permeability of poly(vinyl alcohol) or PVA is surprising, as this is the basis of many commercial membranes in the area (Jonquie`res et al., 2002). The presence of some polarity such as ether or sulphonate groups is desirable. The introduction of sulphonate or carboxylate groups into the PVA structure would seem to be worthwhile. What is absent in the list of polymers tested here for vapour permeation are polysalts formed from cationic and anionic polyelectrolytes, which have good performance in pervaporation dehydration applications. For the removal of water from water/organic liquid or vapour mixtures, hydrophilic organic polymers are generally used because water is readily incorporated and diffused through these materials. The hydrophilicity is caused by groups present in the polymer chain that are able to interact with water molecules. Examples of hydrophilic polymers are: cellulose acetate, ionic polymers, PVA and polyacrylonitrile. Hydrophilic features are found to be essential for good pervaporation performance in the dehydration of ethanol (Bolto et al., 2011).
2.1.
increased by a factor of 2e5, and there is no adverse effect on selectivity (Van Wijk and Jansen, 1990).
2.2.
For the drying of compressed air the preferred membrane material is polydimethylsiloxane or cellulose acetate, because of their acceptable H2O/N2 selectivities and high permeabilities (Sijbesma et al., 2008). Increasing the level of sulphonate groups in poly(ether ether ketone) membranes has been shown to not only increase the permeability of water vapour, but to decrease gas permeability (Liu et al., 2001). The permeability towards water is doubled in poly(aryl ether sulphone) membranes by introducing carboxylate groups, and the H2O/N2 selectivity can be raised to more than 105 (Wang et al., 2001).
2.3.
For the dehydration of natural gas the preferred highly selective membranes are sulphonated poly(ether sulphone) and a (polybutylene terephthalate)/(poly(ethylene oxide) or PBT/PEO block copolymer of molecular weight 1000 Da (Metz et al., 2005). The latter polymer (Fig. 1) had soft rubbery amorphous hydrophilic and hard rigid crystalline hydrophobic segments. It has been found that the water vapour permeability decreases with increasing temperature, which was attributed to the decrease in water solubility on heating (Metz et al., 2002). A cellulose ether composite membrane of high water vapour permeability and selectivity is proposed for the removal of water vapour from pressurized gases or gas mixtures (Ohlrogge et al., 2002). The removal of water vapour from a gas mixture can be accelerated by impregnating a membrane composed of regenerated cellulose with a hygroscopic electrolyte such as lithium bromide. The flux is
C O
O
O
Flue gas dehydration
Appropriate polymers for flue gas dehydration are suggested to be sulphonated poly(ether ether ketone), or SPEEK, and the commercial polymer PEBAX 1074, which contains 45wt% of a polyether block amide (nylon 12) and 55wt% PEO (Sijbesma et al., 2008). Data for the latter material have been included in Table 1. Experiments have been conducted on this polymer using an artificial flue gas containing 11.2 vol% water, the remaining gases being nitrogen, CO2 and oxygen. A continuous removal of 0.6e1 kg/m2 h was achieved over a 150 h test at 50 C. Long term tests on a real, aggressive flue gas stream gave an average water vapour removal rate of 0.2e0.46 kg/ m2 h over 5300 h. Although the flux decreased with time because of the deposition of fly ash dust and gypsum crystals on the membrane surface, performance was judged to be adequate. The product water quality was not sufficient for reuse in the steam cycle, but was satisfactory for use as feed for a demineralised water plant.
Dehydration of natural gas
C
Drying of compressed air
2.4.
Dehydration of ethanol
The production of 99.9% ethanol from 94% ethanol has been achieved by vapour permeation (Sander and Janssen, 1991). Although the membrane material is not revealed, the authors remark on the use of the same materials for vapour permeation and pervaporation of aqueous alcohol, and in an earlier article quote PVA as the membrane of choice for this application (Sander and Soukup, 1988). Vapour permeation was claimed to have the advantages of a lower required membrane area, higher flow rates, and the avoidance of harsh chemical reaction conditions. The volumetric flow rate is about 400 times higher for a vapour than a liquid feed at a given mass flow rate and comparable temperature and pressure conditions. Nevertheless, pervaporation seems to be well established in
O
C
C O
O
O
X Fig. 1 e Structure of PBT/PEO block copolymer (Metz et al., 2005).
O
m Y
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this area. A composite PVA membrane commercialized for pervaporation when tested in vapour permeation mode has yielded fluxes as high as 5.5 kg/m2 h for 90 wt% aqueous ethanol at a pressure of 500 kPa, down to 0.7 kg/m2 h when the pressure was 100 kPa (Jansen et al., 1992). The addition of sodium montmorillonite clay to PVA decreased the water vapour permeation rate from 0.44 to 0.38 kg/m2 h for 90 wt% ethanol when there was 10 wt% clay present (Yeh et al., 2003). This was ascribed to an increase in the tortuosity of the diffusion pathway because of the barrier properties of the dispersed clay. Composite membranes of PVA coated on Nylon-4 had a flux of 0.088 kg/m2 h in the vapour permeation dehydration of 10 wt% aqueous ethanol at 25 C (Lee et al., 1992). In pervaporation mode the flux was 0.42 kg/ m2 h, but the respective separation factors were 94 and 14. Using microporous silica membranes formed by coating a microporous silica layer onto a ceramic porous tube (GallegoLizon et al., 2002), a water flux increase of approximately 100 times to 10 kg/m2 h has been reported for an increase in temperature from 25 to 120 C during the dehydration of a 96 wt % ethanol/water mixture, whilst the ethanol flux remained low at less than 0.1 kg/m2 h (van Veen et al., 2001). The dehydration of aqueous ethanol by vapour permeation has been studied using tubular NaY zeolite membranes at 90e110 C (Sato et al., 2008). For a 90 wt% ethanol feed at 110 C a flux of 20 kg/m2 h was obtained, to give a retentate that was 96.1 wt% ethanol. At a flow rate of 5 kg/m2 h the product was 98.5 wt% ethanol. For a pure water feed a flux of 80 kg/m2 h was achieved. With NaA zeolite membranes and 10 wt% water content a vapour permeation flux of 11 kg/m2 h at 125 C has been obtained (Richter et al., 2006). Further results have been obtained at lower temperatures: 2.5 kg/m2 h at 90 C and 5 kg/ m2 h at 110 C (Sommer and Melin, 2005). Some change in the crystal structure of LTA zeolite has been observed after its use in dehydrating aqueous ethanol (Kyotani et al., 2009). A small deterioration in the water permeance was observed, and some 10% of sodium ions were detached, indicating proton exchange. Membranes composed of aromatic polyamides made from the reaction of bis[4-(4-aminophenoxy) phenyl]diphenylmethane with the diacids terephthalic acid, 5-tert-butylisophthalic acid or 4,40 -hexafluoroisopropylidenedibenzoic acid have been used in vapour permeation experiments (Fan et al., 2002). The highest flux of 0.2 kg/m2 h was obtained with the fluorinated species. It was found that the permeation rate could be increased by introducing a bulky group into the polymer backbone. Other fluorinated polymers have given similar results, with best flux being 0.38 kg/m2 h (Teng et al., 2000). Polyimide membranes have been blended with sulphonated polyether sulphone to give enhanced results of 0.64 kg/m2 h for an air sweep mode of operating vapour permeation at 80e100 C (Wu et al., 2002). Earlier work on a series of polyimide membranes had shown large variations in performance in treating 10 wt% aqueous ethanol at 75 C, the permeation of water being some 10 times greater in the polyimide made from anhydride and dia3,30 ,4,40 -biphenyltetracarboxylic minodiphenylsulphone compared to that prepared from pyromellitic anhydride and 4,40 -oxydianiline (Okamoto et al., 1992). This was attributed to the high rigidity and bulkiness of the polymer backbone in the former case preventing close chain packing and giving a more open structure for water transport.
The vapour permeation performance of symmetrical and asymmetrical polycarbonate membranes, prepared by dryphase and wet-phase inversion methods, has been studied for aqueous ethanol (Wang et al., 2005). In the case of the symmetrical polycarbonate membrane, vapour permeation was found to have a significantly increased separation factor and a slightly decreased permeation rate compared to pervaporation, the respective permeation rates being 0.16 and 0.18 kg/m2 h for a 20 wt% aqueous ethanol at 25 C.
2.5.
Dehydration of isopropanol
The performance of commercially available crosslinked PVA and microporous silica membranes for the dehydration of isopropanol (IPA)/water mixtures by pervaporation/vapour permeation has been studied (Gallego-Lizon et al., 2002). The PVA membranes used in the study belong to a generic family of asymmetric composite membranes for water permeation. These membranes have a supporting layer of nonwoven porous polyester onto which an ultrafiltration membrane is cast and finally a layer of crosslinked PVA. For the range of conditions investigated in the study, water fluxes generally increased with water concentration and increased with the operating temperature from 70 to 105 C. The water flux through the microporous silica membrane was found be to up to three times higher than that through the PVA membrane. A water flux of 21.5 kg/m2 h across the microporous silica membrane was reported at 105 C for a 12.5 wt% water mixture. However, the IPA flux through the microporous silica membrane was much higher than that through the PVA membrane.
2.6.
Dehydration of acetonitrile
In a comparison of vapour permeation and pervaporation for dehydrating aqueous acetonitrile with a silica membrane, a total flux of 3.9 kg/m2 h at 70 C has been measured (Fontalvo et al., 2005). In an economical evaluation it was ascertained that at high water concentrations or just for overcoming the azeotropic composition, vapour permeation was preferred, but for water concentrations lower than the azeotropic amount, pervaporation was the better option.
2.7.
Steam recovery
Water vapour as steam is the most universal energy carrier. Its application is wide spread and can be found in all aspects of industrial processes. Industry converts more than 70% of the fuel it purchases for energy into steam. Waste steam can be found almost in every plant/factory where steam is used, from large industrial establishments such as refineries, power plants, chemical factories, steel makers, ore mining, to medium and small plants such as sugar mills and food processing facilities. It is usually contaminated and cannot be reused. A common practice in dealing with it is to use a condenser to collect water or to discharge the steam to the atmosphere. Discharging the spent steam to the atmosphere is not only an energy loss, but is at the same time an environmental issue as water vapour is a major driver of greenhouse gas-induced climate change (Hoang and Nguyen, 2009).
w a t e r r e s e a r c h 4 6 ( 2 0 1 2 ) 2 5 9 e2 6 6
The steam consumption in a typical thermal power station of a 1000 MW capacity is about 2800 t/h, which translates to about 800 kg/s of condensate. A 10,000 tonne/day discharge of waste steam represents a loss of w0.7 million m3 of natural gas/day or w$40M/year (at 0.4 cent/MJ wholesale to industry). While no direct figures are available in Australia, recycling of steam has significant potential to improve the overall energy efficiency of industry which accounts for 25 Mt CO2/year emissions. With higher energy costs and a growing concern regarding environmental impact, it is highly desirable to recover the energy loss by recycling the spent steam. Membrane processes are being sought that effectively separate contaminants and recover the cleaned industrial waste steam. A high rate of diffusion of water vapour through a nonporous ionic membrane is reported as the key to producing high purity steam at temperatures approaching or exceeding 100 C (Spiegelman and Blethen, 2006). The preferred membrane polymer was a copolymer of perfluorinated ethylene and perfluorinated vinyl compound containing an acid group (sulphonic or carboxylic) or its salt. An example is the sulphonic acid copolymer Nafion, a copolymer of tetrafluoroethylene and perfluoro(4-methyl-3,6-dioxa-7-octene-1-sulphonic acid). A membrane that is substantially gas impermeable is desirable, and this copolymer has a permeability of water vapour that is more than three orders of magnitude greater than the permeability of CO2 or CO, and some six orders of magnitude greater than the permeability of oxygen or nitrogen. PVA/silica nanoparticle composite membranes have been tested for steam recovery (Hoang and Nguyen, 2009). Water vapour fluxes of 70e150 kg/m2 h were achieved at a differential pressure of 6 bars. A steam permeation plant that uses an LTA zeolite membrane has been installed at a sugar works. It treats 93 wt% bio-ethanol, obtaining a permeate that is below 0.1 wt% ethanol (Caro and Noack, 2008). Quite high fluxes are reported: 11.9, 14.9, 17.6 and 22.4 kg/m2 h at 100, 110, 120 and 130 C respectively.
2.8.
Miscellaneous water removal applications
Vapour permeation studies of mixtures propanol-methanolwater and propanol-methanol have been performed with a commercial hydrophilic PVA-polyacrylonitrile composite membrane (Will and Lichtenthaler, 1992). Also, the binary systems ammonia-water and methylamine-water were investigated using a commercial amine resistant PVApolysulphone composite membrane. With the exception of the non-aqueous propanol-methanol system the separation factors and fluxes obtained for the binary systems were sufficiently high for practical application. In the case of the ternary mixture, vapour permeation also showed a much better separation than pervaporation and a reasonable flux. This was not true for the non-aqueous system though, when the flux was extremely small at 98%
Turek et al., 2009 Oren et al., 2010
Secondary RO
Evaporation
80e90%
Secondary RO
Brine concentrator þ Pond
w100%
Mohammadesmaeili et al., 2010a, b Bond and Veerapaneni, 2008
Primary RO
Type B ZLD scheme
Overall recovery
Primary RO Primary RO
Lime-soda treatment Intermediate treatment (several options)
Acevedo et al., 2010
274
w a t e r r e s e a r c h 4 6 ( 2 0 1 2 ) 2 6 7 e2 8 3
Fig. 3 e Bipolar membrane electrodialysis scheme for obtaining mixed acid and basis.
membrane performance. However, the addition of these compounds can have negative effects in the post-treatment of RO concentrates. Greenlee et al. (2011, 2010a, 2010b) studied the effects of anti-scalants in the precipitation of scaling salts, like calcium carbonate, as intermediate stage in a tandem RO system. It was demonstrated that anti-scalants decreased the precipitate particle size and change the shape of the particles. Following the objective of increasing the global recovery of RO operation, several authors (Acevedo et al., 2010; Gabelich et al., 2007, 2011; Ning et al., 2006; Rahardianto et al., 2010) proposed RO tandem processes combined with an intermediate unit in order to eliminate the most problematic substances such as foulants and scalants (Type A scheme in Fig. 2). Acevedo et al. (2010) evaluated the possibility for silica concentration reduction by using ion exchange columns with strong basic anionic resins, and reported that the removal of silica strongly
depends on the pH being available at pH over 8.5, when it also takes place calcium and magnesium precipitation. Therefore, Acevedo proposed as another phase of the research a train of cationic and anionic resins to first remove hardness and allow to increase the pH without precipitation. Studies about salts and silica removal were also developed by Gabelich et al. (2007, 2011) who proposed intermediate chemical demineralization (ICD) that consisted on a solid contact reactor (SCR) followed by a filtration stage. Ning et al. (2006) used a limesoftening unit to reduce the limiting foulants of silica and barium sulphate. With the same objective of eliminating antiscalants and fouling compounds, Rahardianto et al. (2010) proposed a two step chemically-enhanced seeded precipitation (CESP) for accelerated desupersaturation. They reported that the resulting lab and pilot plant data would serve as possible scale-up information although control and operation strategy had to be developed to maintain consistently high demineralization efficiency at industrial level. Other schemes have been investigated with the aim of recovering both salts and water (Neilly et al., 2009). These systems are characterized for having one or more final steps of post-treatment of the concentrate obtained from the second RO stage (Type B scheme in Fig. 2) (Macedonio et al., 2011; Ning and Tarquin, 2010; Oren et al., 2010; Turek et al., 2009; Zhang et al., 2010). Combining WAIV technology with a membrane crystallizer (MCr), Macedonio et al., 2011 achieved recovery factors as high as 88.9% and capital cost of WAIV was reduced in a 64% in comparison to conventional ponds. Ning and Tarquin (2010) studied the fractional crystallization of salts before thermal evaporation of water. In turn, Turek et al. (2009) studied the combination of RO with electrodialysis reversal (EDR) achieving 91.6% of water recovery, although the synthetic water used had no comparable composition with real RO concentrates. Oren et al. (2010) proposed a system based on further concentration of the BWRO brine with EDR and WAIV. Scaling of ED unit was prevented by acidification, operating the electrodialysis in a reversal mode, and a side loop crystallizer which prevented build-up of scaling components. The super-concentrate from the EDR unit was further concentrate in a wind powered WAIV unit that brought final TDS to more of 30%, and showed
Fig. 4 e Experimental set-up for treatment of RO brine using ozone D BAC column D CDI (Lee et al., 2009a, b).
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Fig. 5 e Salt recovery scheme used by Mohammadesmaeili et al. (2010a, b).
promise as a method to recover mineral byproducts. Cost effectiveness of these process alternatives need to be assessed in order to determine their availability and profitability. The third type of ZLD scheme is based on combining type A and type B ZLD scheme. Mohammadesmaeili et al. (2010a, b) proposed a system with and intermediate stage of lime softening and several stages of evaporationecrystallization after the secondary RO (see Fig. 5). Results of this research are explained in more detail in Section 3.2. This type of scheme was also studied by Bond and Veerapaneni (2008); they affirmed that treatment costs and energy requirements of ZLD can be reduced by adding the intermediate steps of concentrate treatment and secondary RO. They evaluated different
options of intermediate treatment: chemical precipitation with sodium hydroxide (NaOH) or lime [Ca(OH)2], fluidized bed crystallization (FBC), adsorption on activated alumina (AA), chemical precipitation with alum [Al2(SO4)$14H2O] and ion exchange. FBC has been demonstrated to produce a much smaller volume of solids than that produced by conventional softening. Besides, the evaluation of treatment costs comparing the system without intermediate treatment (benchmark process) with the system with FBC and microfiltration (ZLD process) shows that treatment cost would be reduced by 50e70%, and energy consumption would be reduced by 60e75%. An important number of brine concentrators are in operation in United States. Operating
Fig. 6 e Electrochlorination system proposed by Badruzzaman et al., 2009.
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experiences of plants using brine concentrators to concentrate reject streams have shown that this is a viable application and the system is highly reliable (Mickley, 2006). Treatment schemes based on ZLD objective are pointed to be a promising option for the treatment of RO concentrates. Initial estimation about treatment costs indicates that these systems would be feasible, but it would be necessary leap from pilot plant to industrial level to safely evaluate the applicability of these systems.
2.2.2.
Emerging technologies for salts recovery
Recovering commercial products is the final challenge to improve the management of RO concentrates whatever their water source is. If valuable substances are recovered from RO concentrates a double objective will be achieved: the reduction of the environmental impact of RO concentrates disposal added to the improvement of the economy of the global treatment process. Through combination of evaporation and crystallization technologies, valuable salts can be recovered. Hajbi et al. (2010) analysed the solubility diagrams of the hexary system Naþ, Kþ, Mg2þ, Ca2þ/Cl, SO2 4 //H2O to define crystallization paths that they used for the experimental design. Their results assessed the potential of recovering salts like NaCl, KCl, CaSO4$2H2O, MgSO4$7H2O, . through the isothermal and isobaric evaporation of the rejected brine. Several studies on the viability of producing salts from RO concentrates of desalination plants managed by Petroleum Development Oman (PDO) were developed by Ahmed et al. (2003). The viability study was focused on the application of the patented SAL-PROC process consisting on the sequential extraction of dissolved elements from inorganic saline waters in the form of valuable chemical products in crystalline, slurry and liquid forms. The potential products that can be recovered would be gypsumemagnesium hydroxide (mixture), magnesium hydroxide, sodium chloride, calcium carbonate, sodium sulfate and calcium chloride. In light of the estimations, processing the 405 106 L/year of the rejected brine generated in the four PDO desalination plants would make possible to produce commercial salts worth US $895,000 in the most optimistic scenario. Although the economic profit is not so high, it is indisputable that the recovery of potential commercial salts is a great option to improve the cost-effectiveness of desalination processes. Regarding the market of inorganic salts, a quantitative assessment of the global market of recovered salts is not available yet because technologies are in an early development stage, but it has been reported that market could hold the produced salts. According to Ahmad and Williams in the United States 45 million tons of salts are produced annually and about 70% of these salts are used by chemical industries (Ahmad and Williams, 2011.). On the other hand Van Houwelingen stated that in the Netherlands the market of recovered salts would not be a problem since all solids generated in fluidized bed crystallization (FBC) processes for the treatment of reverse osmosis concentrates are put to beneficial use by other industries, such as steel production (Bond and Veerapaneni, 2008). Production of mixed acids and mixed bases using Electrodialysis with Bipolar Membranes EDBM (see Fig. 3) is another viable option. Mavrov et al., 1999 obtained suitable acids and
bases mixtures for the regeneration of ion exchangers by processing RO concentrates from desalination of surface water. EDBM has been also applied to RO concentrates from wastewater treatment plants, so this technology will be widely explained in the Section 3.2. Jegatheesan et al. (2009) evaluated the recovery of valuable metals from concentrates by means of liquideliquid extraction. The recovery ratio for rubidium using BAMBP [4-tertbutyl-2-(a-methylbenzyl) phenol] or 4-sec-butyl-2-(a-methylbenzyl) phenol] as extractants was higher than 80% (Jeppesen et al., 2009). Liquideliquid extraction has been also applied by Le Dirach et al. (2005) to recover materials from the concentrated brine of the RO plant in Skhira (Tunisia). The treatment process is divided in several extraction stages: phosphates extraction using a blend of iron and aluminium sulphates, liquideliquid extraction to recover cesium, liquideliquid extraction to recover indium and gallium, rubidium extraction, evaporation of the remaining brine to recover carnallite crystals KMgCl3$6H2O and as final step, extraction of germanium and magnesium. Still, these separation protocols have been studied at viability scale and require final demonstration.
3. RO concentrates from wastewater treatment plants: treatment alternatives In recent years, reverse osmosis has been also applied to further treat the secondary effluents of wastewater treatment plants. These RO concentrates present less salinity than RO concentrates from desalination plants although larger amounts of organic matter, including persistent micropollutants, are contained. Solley et al. (2010) reported that the contaminants in these streams could be 6e7 times more concentrated than in the feed water. Therefore specific treatments, such as advanced oxidation processes (AOPs), have been researched in order to reduce the pollutant load. Most of the reported studies are focused on the reduction of a global parameter that resembles the organic contamination such as the chemical oxygen demand (COD) and the Total Organic Carbon (TOC), although the removal of other pollutants has been also explored. For example, Ersever et al. (2007a, b) studied the removal of nitrogen compounds from RO brines through biological nitrificationedenitrification and sulfate reduction via a fluidized bioactive absorber reactor (FBAR) process and using granular activated carbon (GAC). Experiments were conducted at different hydraulic retention times and nitrate concentrations, showing that the FBAR process could be an efficient technology for nitrogen removal of RO concentrates, although nitrogen was considered a minor pollutant, recent characterisation of RO concentrates from WWTPs shows concentrations of ammonia up to 120 mg/L (see Table 2), turning the interest to the referred technologies. Traditional treatments such as coagulation and activated carbon adsorption were tested for DOC (dissolved organic carbon) removal in RO concentrates (Dialynas et al., 2008). Experimental results showed that ferric chloride is a better coagulant than aluminium sulfate (Al2(SO4)3$18H2O), achieving 52% removal of DOC. Ferric coagulation tests effectively removed colour (79%) while DOC and COD mostly due to low molecular weight compounds were removed up to
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34% and 49% respectively (Bagastyo et al., 2011b). However, other authors (Zhou et al., 2011) reported lower efficiencies of coagulation, attributed to the different characteristics of the wastewater investigated. Adsorption on activated carbon (AC) was studied at different carbon doses. The highest removal observed was 91% for the dose of 5 g/L, so results of DOC removal with GAC adsorption are more promising than in coagulation tests. Similar results have been reported by Zhou et al. (2011), and as in previous works, the organics remaining in the RO concentrate were hardly adsorbed at higher AC dosage, indicating the existence of non-adsorbable organic matter fraction.
3.1. Advanced oxidation processes (AOPs) applied to RO concentrates from WWTPs Table 4 summarizes the research on the application of AOPs to RO concentrates. The studies include the well known ozonation and Fenton technologies, the latest developments in photocatalysis and photooxidation, together with new applications of sonolysis and electrooxidation.
3.1.1.
Ozonation
Feasibility tests of advanced oxidation processes for the removal of organics from the RO concentrate obtained in a water reclamation plant located in Singapore were performed by Zhou et al. (2011). Simple ozone treatment could only remove a small fraction (22%) of DOC with marginal improvements at higher energy input. Nevertheless, O3 treatment provided better performances that UV irradiation or hydrogen peroxide addition. Meanwhile, it was observed that the combination of ozonation with photocatalysis (UVA/ TiO2) enhanced DOC removal up to 52%, while the highest overall efficiency was obtained by the combined UVA/TiO2/O3 with coagulation pretreatment. Moreover, it was found that the latter combination increased the biodegradability index (BDI ¼ BOD5/COD) of the raw RO concentrate by 7e20 times. The O3/H2O2 treatment applied by Westerhoff et al. (2009) achieved a 75% removal of DOC contained in RO concentrate from the Scottsdale Water Campus facility, although very high
ozone (1000 mg/L O3) and hydrogen peroxide (0.7 mol H2O2/ mol O3) dosages were required. Ozone pretreatment for a biological activated carbon unit increased the biodegradability of the RO brine obtained from a domestic wastewater treatment facility (Lee et al., 2009a). Laboratory scale batch experiments showed that ozonation alone achieved a maximum TOC removal rate of 24%, while Biological Activated Carbon (BAC) alone provided a limited 23% of TOC removal. However, the synergistic effect of combining ozone and BAC showed a good potential in reducing fouling problems for a subsequent capacitive deionization (CDI) stage (Lee et al., 2009a, b) (see Fig. 4), a method for removing salts contained in aqueous solutions by electroadsortion (Oren, 2008). A capacitive deionization cell unit consists of two electrodes made of activated carbon and separated by the circulating ion containing solution. Electrons are not transmuted by red-ox reactions but by electrostatic adsorption (Strathmann, 2010). On the basis of the results it was claimed that CDI might be an attractive option for desalting water due to its lower energy demand compared to RO. RO concentrates from wastewater treatment plants contain, among many other pharmaceuticals, beta blockers, that are classified as potentially toxic to aquatic organisms. Because of beta blockers molecules have moieties (amine groups, activated aromatic rings), which are reactive towards ozone, it was tested whether ozonation can be applied for their mitigation. Moderate ozone doses (5e10 mg/L) were found to be sufficient to remove beta blockers efficiently (Benner et al., 2008).
3.1.2.
Fenton process
Fenton process for the treatment of RO concentrates has not been widely studied Westerhoff et al. (2009) developed Fenton (Fe2þ/H2O2) and Fenton-like (Fe3þ/H2O2) experiments in a jartest apparatus. The Fenton process (pHw3.3; 10 mM Fe2þ and 10 mM H2O2) removed up to 50% of DOC; residual iron was precipitated by raising the pH back to 7.5e8.0. Higher DOC removals might have been achieved using higher chemical dosages. No more studies regarding application of Fenton process to RO concentrates have been found in literature.
Table 4 e Advanced oxidation processes (AOPs) applied to RO concentrates. AOP applied
Objective
Photocatalysis
TiO2 þ hv/e þ hþ hþ þ H2 O/HO þ Hþ
Photooxidation
H2 O2 þ hv/2HO
Sonolysis Electrooxidation
H2 OþÞÞÞ/H þ HO 2HO /H2 O2 H2 O þ Anode/Anode½HO ads þ Hþ þ e
Ozonation
O3 /Initiators ðH2 O2 ; UVÞ HO
Fenton8
Fe2þ þ H2 O2 /Fe3þ þ OH þ HO
DOC removal DOC removal with coagulation pretreatment (FeCl3) DOC removal 95% DOC removal DOC removal DOC removal (low) DOC removal COD and TAN removal COD and TAN removal COD, ammonium and emerging pollutants N-nitrosodimethylamine destruction DOC and pharmaceuticals and pesticides oxidation DOC removal Beta blockers molecules removal Improving biodegradability of RO brine DOC removal
Reference Dialynas et al., 2008 Zhou et al., 2011 Westerhoff et al., 2009 Bagastyo et al., 2011a, b Westerhoff et al., 2009 Dialynas et al., 2008 Zhou et al., 2011 Van Hege et al., 2002 Van Hege et al., 2004 Pe´rez et al., 2010 Chaplin et al., 2010 Radjenovic et al., 2011 Westerhoff et al., 2009 Benner et al., 2008 Lee et al., 2009a, b Westerhoff et al., 2009
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3.1.3.
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Photocatalysis and photooxidation
From the various advanced oxidation processes, heterogeneous photocatalysis with TiO2 as catalyst brings the advantage of the possible use of solar radiation. Several recent works deal with the application of UV/TiO2 to the removal of organic load from RO concentrates. For a RO concentrate with an initial DOC concentration of 10.2 mg/L, dark adsorption and UVA light experiments were performed (Dialynas et al., 2008) using two different catalyst doses (0.5 and 1 g/L TiO2). In dark adsorption tests, 30% DOC removal was achieved within 50e60 min at both catalyst concentrations. When UVA light is used during 1 h, the DOC oxidation yield was between 49 and 41% at the high and low catalyst level respectively. The combination of photocatalytic oxidation with coagulation as pretreatment, has been also studied (Zhou et al., 2011). For a raw RO concentrate obtained from a water reclamation plant in Singapore the removal of DOC and COD became rather slow after 1 h of reaction time when using both UVC/TiO2 and UVA/ TiO2 treatment processes. This behaviour was assigned to the nature of the remaining organics that were very recalcitrant to both treatment processes. However, the coagulation pretreatment using ferric chloride (FeCl3), provided the conditions for the gradual removal of DOC and COD by photocatalysis. After a reaction time of 6 h, 95% and 72% of DOC could be removed by UVC/TiO2 and UVA/TiO2. Similarly, Westerhoff et al. (2009) achieved 95% DOC removal at a UV dose of 10.4 kW h/m3, being the removal rate nearly independent of the titanium dioxide dose that was varied between 1 and 5 g/L. When the system UV/H2O2 was used, 40% DOC removal with a UV dose of 11.8 kW h/m3 and H2O2 dose of 10 mM was achieved. Photooxidation (UV/H2O2), has been also tested in RO concentrates from a water treatment plant by Bagastyo et al. (2011a). Complete decolourisation and 50e55% of COD removal was achieved.
3.1.4.
Sonolysis
Organic contaminants can also be degraded by the ultrasonic technique (US) in the presence of a catalyst, acids and nonoxidant gases. The chemical effects of ultrasounds are due to the high temperatures and pressures produced during violent collapse of cavitation bubbles. In water, implosion and fragmentation of the bubble is the centre of energy phenomena: temperature, pressure and electric discharges give rise to H2O sonolysis with production of radical species (OH, H, HOO) and direct destruction of solutes (Hoffmann et al., 1996; Farooq et al., 2008). Results of DOC removal obtained with sonolysis show that this process is less efficient than other AOPs. Testing a RO concentrate with an initial DOC of 10.1 mg/L, only a 29% of the DOC was oxidized in 1 h at 67.5 W. If the energy level is duplicated (135 W) during 1 h of treatment, DOC removal is only increased to 34% (Dialynas et al., 2008). Due to the low DOC removal, the possibility of using sonolysis combined with other AOPs has been studied (Zhou et al., 2011). Using a RO concentrate from a wastewater treatment plant with an initial DOC of 18 mg/L, at 1 h of reaction time, the sequence of DOC removal efficiencies achieved was US < US/H2O2 < US/ O3 < US/H2O2/O3 achieving w31% DOC removal as the highest removal with the last combination.
3.1.5.
Electrochemical oxidation
Electrochemical oxidation is a very efficient alternative for the treatment of wastewater containing non-biodegradable organics and ammonium compounds. With regard to RO brine, electrochemical treatment seems a promising technology, as the high salinity of the RO concentrate ensures an excellent electric conductivity that could reduce the energy consumption. Moreover, the high chloride content could facilitate indirect bulk oxidation through the electrogenerated strong oxidants such as hypochlorite and oxidation of total ammonia nitrogen (TAN) and organics can be accomplished simultaneously (Van Hege et al., 2002; Pe´rez et al., 2010). Reverse osmosis concentrates from a pilot installation processing a mixed domestic and textile wastewater effluent were tested to determine the characteristics and efficiency of several anode materials (Van Hege et al., 2002, 2004). Comparing RuO2 and Boron Doped Diamond (BDD) anodes, results showed that current efficiency for BDD anode (35.2%) is more favourable than the efficiency obtained for a RuO2 anode (14.5%). The higher efficiency of BDD anodes can be attributed to the superior chlorine production rate of this material because of the high selectivity towards the chlorine evolution reaction. Longer treatment times showed the capacity of BDD anodes to provide total removal of ammonium and DOC contained in the RO concentrate generated in a reverse osmosis facility aimed at the industrial reuse of the reclaimed water (Pe´rez et al., 2010). Thanks to the good results obtained in COD and ammonia removal, electrochemical oxidation of RO retentates with BDD anodes was also studied for the removal of emerging micropollutants (Pe´rez et al., 2010) Membrane concentrates contain an increasing amount of salts, organics and biological constituents. Among those organic compounds appear the socalled “emerging pollutants” which include pharmaceuticals, personal care products and other metabolites. These compounds appear in RO concentrates because they are highly rejected by the reverse osmosis membranes, and their removal must be investigated because of the environmental risk associated to their emission to the receiving natural water bodies. RO concentrates from a UF/RO pilot plant that treats the secondary effluent of a WWTP were oxidised electrochemically using BDD anodes (Pe´rez et al., 2010). Ten emerging pollutants that were found as most prevalent in the secondary effluent of a WWTP were investigated achieving removal percentages higher than 92% after 2 h of electrooxidation. Ibuprofen appears like the most resistant compound to electrochemical treatment, thus needing longer times for complete removal. The influence of the applied current density and the initial concentration of the micropollutants, obtained after different recovery ratios of the RO operation, allowed to conclude that the controlling step of the process kinetics is the mass transfer from the liquid bulk to the surface of the BDD anode due to the low concentration of the target compounds. Moreover, the formation of organochloride compounds was kept at low levels. However, testing RuO2/IrO2-coated Ti electrodes for elimination of pharmaceuticals and pesticides, Radjenovic et al. (2011) concluded that this type of electrodes should not be used, because chlorine mediated indirect oxidation was the main
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mechanism and the organic compounds were transformed into their halogenated derivatives, which are persistent and require extreme treatment conditions. Advantages of BDD anodes were also probed for the elimination of N-nitrosodimethylamine, a highly toxic disinfectant compound present in chlorinated wastewater effluents, and further concentrated in RO retentates (Chaplin et al., 2010). Research focused on the electrooxidation of this compound using BDD electrodes showed its advantage over other AOPs, as oxidation was carried out in the presence of high HCO 3 concentrations. This is an advantage over other advanced oxidation processes due to the fact that scavenging effect of HCO 3 over hydroxyl radical is not a big drawback in electrooxidation.
3.2. Recovery of nutrients and salts from RO concentrates from WWTP’s Evaporation and crystallization steps have been evaluated by Mohammadesmaeili et al. (2010a, b) to recover salts from RO concentrates. To maximize the salts recovery the authors proposed a combined system based on a RO tandem with an intermediate stage of lime softening, and several stages of evaporation-crystallization after the secondary RO system (see Fig. 5). Thanks to the lime-soda treatment, magnesium hydroxide with a purity of 51e58% was obtained. This precipitate also contained 19.5e23.3% of CaCO3 and 1.3e7.8% of CaSO4 as impurities. Besides, calcite (CaCO3) with a purity of 95% and calcium sulfate with a purity of 92% were obtained as byproducts (Mohammadesmaeili et al., 2010a). After this softening treatment, the brine was treated in a second reverse osmosis unit. Then, an evaporation and crystallization step was carried out and a mixture of Na2SO4 (86e88%) and NaCl (5e14%) was obtained. The final step of the treatment train is an evaporation to dryness of the left effluent and as a result, þ a mixture of Naþ, SO2 4 , Cl and K (35e40%, 17e33%, 25e41%, 1e5% respectively was obtained) (Mohammadesmaeili et al., 2010b). The combination of reverse osmosis with evaporation and crystallization has been applied at the Doswell Combined Cycle Power Plant in Hanover County (Virginia) (Seigworth et al., 1995) with the goal of achieving zero liquid discharge. However, in this case the filter cake from the press must be disposed off. Thus, salts recovery by combining evaporation and crystallization is highlighted as a feasible option, although further work is still needed to evaluate the economic aspects of salt production. Zhang et al. (2009) analysed the viability of recovering valuable compounds from RO concentrates by applying ED. They used ion exchange membranes for the separation of nutrient ions and organic compounds from salts contained in the RO concentrates coming from the treatment of wastewaters of a food company. RO concentrates contained 120 mg/L of phosphate and 120 mg/L of organic compounds measured as TOC. The experimental tests were carried out with synthetic solutions resembling the composition of this RO concentrate (0.45 g/L NaCl, 0.54 g/L MgSO4, 0.026 g/L NaNO3, 0.69 NaHCO3, 0.062 g/L Na2HPO4 and 0e120 g/L TOC). ED tests were carried out to evaluate the performance of two membranes: a non-selective ED membrane (SA) and an ED membrane selective for monovalent anions (MVA). Experimental results reflected that using the MVA membrane,
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monovalent anions can be separated from multivalent ions, that later are retained in the diluate compartment. The separation factor was improved by increasing the pH or reducing the applied current. With regard to the organic compounds it was concluded that larger ions are retained more efficiently than the smaller ones, which points to a sizeexclusion mechanism, while zwitterions are retained almost completely in the diluate side. In the same study it was affirmed that in a test carried out with real RO concentrates, more than 85% of the organic fraction remained in the feed while salt concentrations (Cl and SO2 4 ) decreased, suggesting that the separation of salts from organics by electrodialysis is feasible. Later studies (Zhang et al., 2010) endorsed that high overall recovery using ED is achieved and a moderate selectivity was observed between SO2 4 /Cl and HCO3 /Cl and 2þ a moderate to high selectivity was found between Ca /Naþ by using the non-selective anion and cation exchange membranes. Production of mixed acids and mixed basis using Electrodialysis with Bipolar Membranes (EDBM) (see Fig. 3) is also a viable option. For conversion of RO concentrates from wastewater treatment plants into mixed acid and mixed base streams, Badruzzaman et al. (2009) evaluated the use of bipolar membrane electrodialysis (EDBM). Moreover, they proposed electrochlorination (see Fig. 6) for onsite chlorine generation. RO concentrates contain a significant amount of divalent metals, so a Naþ based cation exchange resin was used to remove them in order to avoid potential precipitation and scaling on the membrane surface. Acid and base production tests were conducted in a batch configuration. The salt concentration in the feed in terms of conductivity went down from 9 mS cm1 to less than 2 mS cm1, while the acid and base concentrations of the products was around 0.2 N. Regarding the separation of nutrients from RO concentrates, adsorption using polymeric resins has been investigated to recover phosphate. The technology seems feasible although phosphate concentrations are not very high, Kumar et al. (2007) analysed the separation of phosphates using polymeric ligand exchange (PLEs) resins and their recovery as struvite. RO concentrates were collected from two pilot integrated membrane systems at the wastewater treatment plant in Rio Rancho (New Mexico) and the North City water reclamation plant in San Diego. The RO concentrate with a low chloride concentration (0.684 g/L Cl) and containing 10 mg/L of phosphate was passed through a packed column loaded with a PLE resin. The regeneration of the exhausted column was conducted using NaCl. In the regenerant solution, P was recovered at a concentration of 78 mg/L. To achieve struvite precipitation, tests were conducted with deionized water spiked with 20 mg P/L to determine the best molar ratio (P:Mg2þ:NHþ 4 ::1:1.5:1) and then with the recovered solution. Phosphate can be potentially recovered as struvite, although nitrate and sulfate also precipitate.
4. RO concentrates from industrial water sources: treatment alternatives Hypersaline brine production is one of the most important environmental issues in mining industry. RO concentrates
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have a conductivity similar to RO concentrates from desalination plants (22,000 mS cm1) and regarding anion composition, sulfate concentration is much higher than chloride (16,000 mg/L SO2 4 ; 955 mg/L Cl ) (Randall et al., 2011). According to this composition, crystallization techniques based on eutectic freeze diagrams were investigated. Eutectic freeze crystallization (EFC) is based on achieving the eutectic temperature as means to separate aqueous solutions into pure water and solid salts. Because the heat of fusion of ice (6.01 kJ/mol) is six times less than the heat of evaporation of water (40.65 kJ/mol), the energy required to separate the water as ice is significantly lower than that required to separate it by evaporation. In addition, the simultaneous production of pure ice and pure salt(s) is a major advantage. The economic advantage of EFC has been analyzed by Nathoo et al. (2009). The authors compared EFC in a tandem crystallizer (see Fig. 7) with an Evaporative Crystallization (EC) system to obtain NaCl$2H2O and Na2SO4$10H2O. Tests were carried out using reverse osmosis retentates from typical mine water brines containing high levels of sodium, sulfate and chlorine. The temperature in each crystallizer was selected according to the concentrates composition in order to obtain the maximum salt recovery. Taking into account the electricity cost and the salts production, a comparison between EFC and EC concluded that cost saving of using EFC is in the range of 80e85% for the brines under study. In addition, the revenue generated by the sale of the pure salts produced by EFC and the cost of disposal of the mixed salt waste produced by EC were not included in the cost evaluation developed in that study. The EFC technique has been also evaluated for the treatment of multicomponent hypersaline brine form the eMalahleni (South Africa) Water Reclamation Plant (EWRP), built to address the damage done to the environment and water systems through Acid Mine Drainage. It was concluded that the EFC process could increase the production of the water reclamation plant by 120 m3/d while a production of 476 kg/ d CaSO4$2H2O and 471 kg/d Na2SO4$10H2O would also be obtained (Randall et al., 2011). The recovery of pure salts is difficult because they usually have several hydrate forms. Sodium sulfate exists in three hydrate forms: anhydrous Na2SO4 (thenardite), Na2SO4$7H2O and Na2SO4$10H2O (mirabilite). Reddy et al. (2010) study the recovery and purity of sodium sulfate using synthetic
Fig. 7 e Simplified EFC process flow sheet (Nathoo et al., 2009).
solutions with a composition simulating brine from the mining industry. Pure water crystals were obtained (90% pure Na2SO4 crystals was obtained from a concentrated NaCl stream prior to any sodium chloride crystals being produced.
5.
Conclusions
Since direct disposal of RO concentrates from desalination plants is recognised as a practice with adverse impacts on marine ecosystems, the search for environmentally friendly management options is a technological challenge. Thus, in this review, treatment technologies of water reverse osmosis concentrates have been addressed, classified according to the source of RO concentrates and the level of development of each management technology. As the composition of the concentrate is closely related to its source, and the selection of the most suitable treatment is based on the concentrate composition, three different sources of RO concentrates have been reviewed i) desalination plants, ii) tertiary processes in WWTP and, iii) mining industries. With regard to RO concentrates from desalination plants traditional technologies, such as solar evaporation easy to operate but with a large requirement of land area, give way to more efficient emerging technologies that allow for considerable brine volume reductions such as electrodialysis, forward osmosis, membrane distillation or even its coupling with crystallisation, MDC, that facilitates easy recovery of high purity NaCl crystals. Recovery of salts, nutrients and valuable compounds from RO concentrates is a promising investigation line and thus related works have been collected. However, although thermal brine concentration technologies are well established, they are energy intensive and not cost effective for large-scale applications. In case of freezing processes the initial investment and complexity of operation limit their use. With regard to the treatment of RO concentrates from WWTPs studies are mainly focused on the reduction of pollutant load and removal of micro-pollutants that could accumulate and cause a high environmental impact. Good results have been obtained applying advanced oxidation technologies and among them electrochemical oxidation with BDD anodes for the removal of micro-pollutants and persistent compounds. Advanced oxidation processes are probably the most promising technologies to degrade and detoxify endocrine disrupting compounds, but the high cost of these technologies may limit their application. Alternatives aiming at zero liquid discharge, through the combination of individual technologies, are highlighted as the most promising management option in this investigated field. Although studies addressed to this objective are mostly related to RO concentrates from desalination plants, ZLD systems can be applied to any RO concentrate, regardless its origin source. The final objective of ZLD systems must be the recovery of valuable compounds from RO concentrates. Joining both objectives: ZLD and valuable compounds recovery, the problem of RO concentrates management would
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be solved. However, the main drawback of ZLD systems is the feasibility of application to industrial level due to their investing and operating costs. Although first economic estimations are encouraging, the profitability of these systems has been demonstrated only at lab and pilot plant scale.
Acknowledgements Financial support from projects Consolider CSD2006-44, CTQ2008-00690/PPQ, ENE2010-15585, CTQ2008-03225/PPQ is gratefully acknowledged.
references
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Available online at www.sciencedirect.com
journal homepage: www.elsevier.com/locate/watres
Similarities in effluent organic matter characteristics from Connecticut wastewater treatment plants Matthew L. Quaranta, Mykel D. Mendes, Allison A. MacKay* University of Connecticut, Environmental Engineering Program, 261 Glenbrook Rd Storrs, CT 06269, USA
article info
abstract
Article history:
Effluent organic matter (EfOM) from five Connecticut (USA) municipal wastewater treat-
Received 13 August 2011
ment plants was isolated with DAX8 (hydrophobic fraction) and XAD4 (transphilic fraction)
Received in revised form
resins. Isolate recoveries ranged from 18 to 42% of the total organic carbon for DAX8 resin
7 October 2011
and from 6 to 12% for XAD4 resin. Isolated EfOM was characterized by traditional organic
Accepted 11 October 2011
geochemistry techniques. Weight-averaged molecular weights of extracted EfOM by size
Available online 28 October 2011
exclusion chromatography were 450e670 Da with higher weights observed for the hydrophobic fractions than the transphilic fractions. Fluorescence characterization showed both
Keywords:
humic- and fulvic-like fluorescence, as well as tryptophan- and tyrosine-like fluorescence,
Excitationeemission matrix
the latter not commonly observed for terrestrial organic matter. Fluorescence indices were
NOM
between 1.5 and 1.9 with lower values observed for hydrophobic EfOM fractions than for
SUVA
transphilic fractions. Specific ultraviolet absorbance was measured between 0.8 and
E2/E3 ratio
3.0 L mg1 m1 with higher values for the hydrophobic EfOM fractions. Together these results indicated that isolated EfOM is similar in characteristics to microbially derived organic matter from natural aquatic systems. Little variation in EfOM characteristics was observed between the five wastewater treatment plants, suggesting that the characteristics of EfOM are similar, regardless of treatment plant design. ª 2011 Elsevier Ltd. All rights reserved.
1.
Introduction
Effluent discharges of treated municipal wastewater represent an important source of organic matter to aquatic ecosystems. In 2008, the estimated volume of wastewater treated was 32,345 million gallons per day in the US (EPA, 2008). Treated wastewater effluent typically contains dissolved organic matter (DOM) concentrations that range between 4 and 20 mgC/L, comparable to DOM concentrations in natural water sources (Thurman, 1985). Such effluent discharges can constitute a large fraction of streamflows in systems receiving treated wastewater. For example, nearly a quarter of all US National Pollutant Discharge Elimination System (NPDES) permitted effluent releases have in-stream dilutions of less than 10-fold under average flow conditions
(Brooks et al., 2006). The contribution of organic matter from wastewater effluent discharge to rivers and streams is particularly significant during low flow summer periods and in arid regions where wastewater effluent can contribute up to 100% of streamflow during dry months (Swayne et al., 1980). Little work has examined anthropogenic sources of organic matter, such as in wastewater effluent; thus, it is unknown how effluent organic matter (EfOM) contributes to ecosystem processes in streams receiving wastewater effluent discharges. The importance of dissolved organic matter from natural sources in stream systems is well established. DOM is an important carbon source in stream systems, supporting microbial growth as the base of the food web (Allan and Castillo, 2007). The presence of DOM affects the availability
* Corresponding author. Tel.: þ1 860 486 2450; fax: þ1 860 486 2298. E-mail address:
[email protected] (A.A. MacKay). 0043-1354/$ e see front matter ª 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2011.10.010
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of dissolved metals and organic contaminants by serving as a sorptive phase that can be transported through aquatic environments (Croue et al., 2003; Haitzer et al., 2003; Pernetcoudrier et al., 2008; Polubesova et al., 2008). DOM is also an important mediator of indirect photolysis reactions that are important to the degradation of many organic contaminants, including pesticides, pharmaceuticals, and personal care products (Doll and Frimmel, 2003; Guerard et al., 2009; Lim et al., 2008). Efforts to probe more deeply the role of natural DOM in these ecosystem processes have emphasized bulk physical and chemical characteristics of these materials. Consequently, organic matter from natural systems has been extensively characterized in previous work (Thurman, 1985). The characterization of natural DOM isolated from many sources has indicated two end members that bound DOM characteristics for most aquatic systems (McKnight et al., 2001). These conclusions were drawn by isolating operationally defined humic and fulvic acid components using hydrophobic resin (Thurman and Malcolm, 1981), followed by characterization techniques, including size exclusion chromatography (Chin et al., 1994), ultravioletevisible light absorption- (Weishaar et al., 2003), fluorescence- (Cory et al., 2010; Hudson et al., 2007; McKnight et al., 2001), and nuclear magnetic resonance (Woods et al., 2010) spectroscopy. Findings from these studies highlighted DOM extracts from the Suwannee River, originating as allochthonous DOM from terrestrial sources, and from perennially ice-covered Antarctic lakes, originating as autochthonous DOM from microbial sources, to be end members, between which the characteristics of DOM from other sources fell (McKnight et al., 2001). In general, DOM of terrestrial origin, in contrast to DOM of microbial origin, has higher average molecular weight (Chin et al., 1994), higher specific UV absorbance (Weishaar et al., 2003), a higher fraction of aromatic carbon (Thurman, 1985), but lower fluorescence indices (McKnight et al., 2001). Importantly, these characteristics have been correlated with contaminant fate processes and thus, this continuum model with Suwannee River and Antarctic lake end members also applies to DOM contaminant sorption coefficients (Chin et al., 1997) and contaminant degradation rates (Guerard et al., 2009). In contrast to natural DOM, wastewater effluent organic matter (EfOM) has not been extensively characterized. One challenge to obtaining a comprehensive understanding of EfOM characteristics from existing studies is the lack of standard protocols between studies. For example, different resins have been used to extract DOM from effluent sources, including XAD4, anion and cation exchange resins, and some studies have used organic solvents as eluents (Baker, 2001; Imai et al., 2002; Leenheer, 1981; Pernet-coudrier et al., 2008; Wang et al., 2009). Often only a subset of typical natural DOM characterization techniques are applied to a given EfOM sample, making it difficult to determine whether EfOM characteristics have the same relative trends among characteristics as for natural DOM. Among the available data, EfOM appears be more similar to organic matter derived from natural microbial sources, despite the difference in microbial communities between wastewater treatment plants (predominantly bacteria and archaea) and natural aquatic systems (with a large algal contribution). EfOM appears to have lower molecular weight and lower SUVA than
285
terrestrially derived organic matter and fluorescence excitationeemission matrices that are more similar to microbial organic matter (Baker, 2001; Hudson et al., 2007; Imai et al., 2002; Park et al., 2010). However, on the basis of these few studies, it is unclear what similarities there are between EfOM characteristics from different treatment plants (Chen et al., 2003; Imai et al., 2002), and thus, where EfOM may group relative to the established natural DOM continuum. The purpose of this study was to characterize EfOM from five Connecticut (USA) municipal wastewater treatment plants. Our goals were to compare EfOM characteristics to natural DOM characteristics and to begin establishing the ranges in EfOM characteristic values between treatment plants. To this end, we used widely available characterization techniques e resin isolation, size exclusion chromatography and optical e and fluorescence spectroscopy e because extensive published datasets of NOM characteristics by these methods are available. We sampled EfOM from conventionally operated large (primary and secondary clarification) and small (no primary clarification) activated sludge plants, typical of designs throughout the US. Several plants had unique operating designs (advanced nitrogen removal, UV disinfection, powdered activated carbon) that could contribute to differing EfOM characteristics. This study examining the range of characteristics of EfOM allowed us to establish how EfOM relates to the natural organic matter continuum.
2.
Experimental methods
2.1.
Materials
Methanol (Certified ACS), acetonitrile (HPLC Grade), HCl (Trace Metal Grade), KOH (certified ACS) and acetone (Optima) were purchased from Fisher (Fair Lawn, NJ). DAX 8 Supelite Resin was purchased from Supelco (Bellafonte, PA), XAD4 Amberlite resin and cation exchange resin (DOWEX Marathon MSC, Hþform) were purchased from Sigma Aldrich. Glass Chromaflex chromatography columns with Teflon end fittings and 0.20 mm bed supports were purchased from Kontes (Vineland, NJ). Potassium hydrogen phthalate was purchased from Ricca Chemical Company (Pokomoke City, MD). Polystyrene sulfonic acids (PSS) in sodium salt form were purchased from Polysciences (Warrington, PA). All water was obtained from a high-purity water system (18 MU-cm, Diamond NANOpure, low TOC system, Barnstead-Thermolyne, Dubuque, IA).
2.2.
Treatment plants
Wastewater effluent was collected from five wastewater treatment plants in the state of Connecticut (USA) between November 2010 and April 2011 (Table 1). The sources of raw wastewater to all of the treatment plants consisted primarily of residential wastewater with some small contributions from light commercial businesses. CT1 was configured to receive stormwater inputs; however, we sampled effluent from this plant during dry weather conditions. CT1 is a mid-sized conventional treatment plant with primary and secondary clarification and aerated activated sludge tanks, as typical of many US treatment plants. CT2 is typical of many small-sized
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Table 1 e Overview of wastewater treatment plant operation. Treatment plant CT1
CT2
Influent sources
Ave flow rate
Primary clarification
Mainly residential, commercial office buildings Residential communities
60 MGD
Yes
0.4 MGD
No
CT3
Mainly residential, commercial office buildings
24 MGD
Yes
CT4
Mainly residential, some research laboratories
1.5 MGD
No
CT5
Mainly residential, commercial office buildings
3 MGD
Yes
treatment plants in the US, operating with aerated activated sludge, but no primary clarification. Some nitrogen removal is likely achieved in this plant because the aeration tanks are operated with lower than average dissolved oxygen levels. CT3 is representative of a mid-sized conventional treatment plant that practices nitrogen removal by carbon addition through the modified LudzackeEttinger process. Treated wastewater from CT3 is disinfected with UV light before discharge. CT4 also is typical of a small-scale conventional treatment plant with no primary clarification; however, the EIMCO carousel aeration tank design enables advanced nitrogen removal. CT5, a smallsized treatment plant that utilizes both primary and secondary clarification, is operated with the addition of powdered activated carbon to the aeration tanks that provides media for attached growth. With the exception of CT3, no other plants were operating disinfection treatments prior to effluent discharge at the time of our sample collection.
2.3.
Effluent collection and storage
Wastewater was collected as a grab sample at the discharge point of the treatment plant following disinfection (CT4, other plants were not operating seasonal disinfection). Effluent samples were transported to the lab in opaque 50 L Nalgene carboys, filtered using Whatman GF/A (Maidstone, England) glass fiber filters (1.6 mm), acidified to pH 2 using HCl, and stored in the refrigerator at 2 C within 24 h. Samples were isolated on the Amberlite resins within one week. The isolated fractions were stored acidified (pH 2) and at 2 C, and characterized within four weeks. Between sampling rounds, the carboys were washed with 3 consecutive high-purity water washes to remove residual water from the previous sample.
2.4.
Organic matter isolation
The resin cleaning and organic matter isolation methods using non-ionic DAX 8 and cation exchange resins were adopted from previous studies (Thurman and Malcolm, 1981), with the following changes: organic matter was only extracted
Treatment technology Traditional activated sludge treatment with coarse bubble diffusers Traditional activated sludge treatment with mechanical aeration Activated sludge and modified LudzackeEttinger process, fine bubble diffusers in aeration tanks Activated sludge with EIMCO carousel aeration tanks, mechanical aeration Activated sludge treatment with granular activated carbon suspended growth media
Denitrification
Disinfection
No designed denitrification process Nitrification and denitrification due to low DO in aeration tanks Nitrification and denitrification by modified Ludzacke Ettinger process Nitrification and denitrification by aerobic and anaerobic zones of carousel No designed denitrification process
None
None
UV
None
None
once on non-ionic resins, humic acids were not precipitated from the extracted eluent, and extracted eluent was analyzed directly without lyphilization. In addition, an XAD4 resin column was used to capture a portion of the less hydrophobic fraction of organic matter that is not retained on the DAX 8 resin. The XAD4 resin was prepared in the same manner as the DAX 8 resin. The DAX8 and XAD4 resin columns were connected in series using Teflon tubing. Effluent samples were pumped through the columns at a rate of 15 bed volumes per hour. Approximately 50 bed volumes of each effluent sample were extracted to ensure that a significant amount of organic matter was sorbed. The organic matter was back-eluted from each column individually using 0.1 M potassium hydroxide. Organic matter extracts were deionized to remove cations by pumping the eluted extracts through a strong Hþ cation exchange resin column. The organic matter eluted from the DAX8 resin is the hydrophobic (HPO) fraction, and the organic matter from the XAD4 resin is the transphilic (TPI) fraction. The organic matter that passed through both columns constituted the hydrophilic (HPI) fraction. The deionized, pH 2 EfOM extracts were used for the subsequent analyses. All glassware were cleaned using Liquinox laboratory soap, followed by repeated rinses with high-purity water.
2.5.
Dissolved organic carbon analysis
Dissolved organic carbon (DOC) was measured with a Tekmar Apollo 9000 combustion organic carbon analyzer. Carbon concentrations in the collected wastewater effluent samples were measured after the sample was filtered and acidified. Carbon concentrations of the HPO and TPI fractions were measured after cation exchange resin treatment. A homogenized sample of the HPI fraction was collected after the entire effluent sample had been extracted. The percentage of organic matter in each fraction was calculated on a per mass basis (Table 2). The total percent recovery of the CT3 treatment plant was much lower than the other four plants, possibly because the HPI fraction sub-sample for DOC analysis was not representative.
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Table 2 e Initial effluent organic carbon concentration and mass percentage of effluent organic matter in resin fractions. Location
Effluent DOC (mgC/L)
%Hydrophobic (HPO)
%Transphilic (TPI)
%Hydrophilic (HPI)
6.22 5.57 5.29 4.2 5.93
30.4 19 19.3 42.3 18.8
9.9 11.2 9.6 11.9 6.5
59.6a 79 40.3 57.1 80
CT1 CT2 CT3 CT4 CT5
% Recovered
109.2 69.2 111.3 105.3
a Calculated by difference.
2.6.
Fluorescence and ultravioletevisible spectroscopy
2.8.2.
Excitationeemission matrices (EEMS) were obtained for each sample using a Cary Eclipse (Australia) fluorescence spectrometer with a xenon flash lamp, slit widths of 5 nm and a scan rate of 1200 scans per second. The excitation wavelength was scanned from 200 to 450 nm in 10-nm increments, and the emission wavelength from 250 to 550 nm in 2-nm increments. Samples were analyzed on the fluorometer directly following isolation without dilution or pH adjustment (pH w2) (McKnight et al., 2001). Absorbance spectra for all isolated EfOM samples were obtained at wavelengths between 200 and 550 nm using a Cary 100 spectrophotometer.
2.7.
High pressure size exclusion chromatography
Extracted organic matter molecular weights were obtained using size exclusion chromatography on 50 mL aliquots (Hewlett Packard 1050 series HPLC). The mobile phase was 0.002 M, pH 6.8 phosphate buffer with 0.1 M NaCl (Cabaniss et al., 2000; Chin et al., 1994). A Protein Pak 125 column and matched guard column were purchased from Waters (Milford, MA). Five PSS standards (1 KDa, 1.8 KDa, 4.6 KDa, 18 KDa, 35 KDa) and acetone were used as calibration standards. The diode array detector was set to analyze the PSS standards and acetone at 224 nm and set to 254 nm for organic matter samples.
2.8.
Data analysis
2.8.1.
Organic matter molecular weight
The average molecular weights of the organic matter were determined from size exclusion chromatograms using the following two equations for number-averaged (Mn), and weight-averaged (Mw) molecular weight (Cabaniss et al., 2000; Chin et al., 1994): X N N X Mn ¼ hi ðhi =Mi Þ (1) i¼1
Mw ¼
N X i¼1
i1
X N hi ðMi Þ hi
(2)
i¼1
where hi is the height of the peak corresponding to the compound molecular weight Mi. To ensure that these equations gave accurate values of the average molecular weights: (1) the baseline was corrected by subtraction of a constant intensity across the entire chromatogram, and (2) the low molecular weight cutoff was set to 50 Da to minimize sensitivity to chromatogram tailing at low molecular weights (Cabaniss et al., 2000). The high molecular weight cutoff was set to 3200 Da for all samples, based upon the range of molecular weights observed among our sample set.
EEMS correction
Several corrections of EEMS data are necessary to obtain fluorescence spectra that are inter-comparable among datasets. Excitation intensity corrections were not necessary because our instrument was equipped with a red photomultiplier tube that provides excitation spectra that are not very distorted (Lakowicz, 1999). An emission intensity corrections array was made by comparing a solution of 103 M quinine sulfate solution in 0.05 M sulfuric acid that was scanned daily (Ex 346.5 nm and Em 384e667 by 1 nm) and compared with tabulated data (Lakowicz, 1999) Differences in intensity between our instrument and tabulated values were used to create a wavelength intensity correction matrix that accounted for changes in lamp emission intensity over time. Since excitation intensity corrections were not needed, the wavelength intensity matrix was the array of emission corrections multiplied by an array of ones. EEMS were corrected for the inner filter effect, according to (McKnight et al., 2001; Murphy et al., 2010): Fcorr ¼
Fmeas 10ðAexit þAemit Þ
(3)
where Fcorr and Fmeas are the corrected and measured fluorescences, respectively, and Aexit and Aemit are the sample absorbances at the excitation and emission wavelengths, respectively. The inner filter correction was applied to all fluorescence measurements in each EEM. Raman scatter was removed from fluorescence spectra by using a high-purity water blank analyzed using the same parameters as the samples. The blank was multiplied by the wavelength intensity correction matrix, and the resulting Raman scatter matrix was subtracted from each sample EEM. The subtraction of a water blank was not sufficient to remove scattered light when the excitation and emission wavelengths were equal and when the emission wavelength was twice the excitation wavelength. These light scattering interferences were removed manually by inserting a value of zero for intensity into an EEM when the excitation and emission wavelengths were within 8 nm of each other, and when the excitation wavelength multiplied by 2 was within 8 mm of the emission wavelength. Zero intensity values were also inserted into the EEM when the excitation wavelength was longer than the emission wavelength. Lastly, the EEM was normalized to Raman Units to remove any instrument-specific artifacts by dividing by the area under the Raman peak of the water blank at excitation 350 nm and emission 381e426 nm. Commonly, EEMS are measured with an excitation wavelength range from 200 nm to 450 nm; however, it was necessary to truncate data below 230 nm. At the concentrations of
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organic matter that were used in this study, the absorbance of light below 230 nm was above 1, indicating that less than 1% of light was transmitted through the sample. This high absorbance introduced a very large inner filter correction factor. Combining this large correction factor with the large amount of fluorometer noise at low wavelengths caused large noise peaks in the EEMS. A cutoff at 230 nm allowed the B, T1, and A peaks to be measured without excessive noise.
2.8.3.
Excitationeemission matrix analysis
The Fluorescence Regional Integration (FRI) method was used to quantitatively compare EEMS spectra among samples (Chen et al., 2003). The FRI method separates an EEM into five regions based on observed fluorescence peaks A, B, C, and the two T peaks (for example designation, see Fig. 2). The regions were separated with a horizontal line at 250 nm and two vertical lines at 330 and 380 nm. A diagonal line intersects all the points where emission wavelength equals excitation wavelength. The volume, Fi, under the EEM in each region was calculated using the following equation for discrete measurements (Chen et al., 2003): Fi ¼
XX ex
Iðlex ; lem ÞDlex Dlem
(4)
em
where I is the intensity at any excitationeemission pair, Dlex is the excitation wavelength interval (10 nm) and Dlem is the emission wavelength interval (2 nm) between observations. Since each of the five regions had a different area, the total volume of each region was normalized with the use of a multiplication factor given as the inverse of the area of that region divided by the total area (Table S3, Supplemental Information). The resulting values were average fluorescence response per unit area in each region. Lastly, the average fluorescence response of each region was converted to a percent fluorescence response.
2.8.4.
Optical methods
Three optical methods were used to characterize EfOM. Specific ultraviolet absorbance (SUVA, L mg1 m1) was calculated using the following formula (Weishaar et al., 2003): SUVA ¼
100 A254 DOC
(5)
where A254 is the sample absorbance measured at 254 nm measured in a 1-cm cuvette, and DOC is measured in mgC/L. Fluorescence index was calculated by dividing the fluorescence intensity measured at excitation 370 nm and emission 450 nm by the intensity measured at excitation 370 nm and emission 500 nm (McKnight et al., 2001). The E2/E3 ratio was calculated by dividing the sample light absorbance at 254 nm by the absorbance at 365 nm (Dalrymple et al., 2010).
3.
Results and discussion
3.1.
Effluent dissolved organic carbon
The percent of effluent organic matter (EfOM) isolated in the hydrophobic (HPO) fraction of the wastewater samples was less than typically found for natural organic matter isolates.
All of our effluent samples had similar concentrations of dissolved organic carbon (DOC), ranging from 4 to 6 mgC/L (Table 2). Between 18 and 42% of EfOM was extracted as the HPO fraction (Table 2). These fractions of HPO organic matter are considerably less than the 50e80% of natural DOM from rivers and lakes that is recovered as the HPO fraction (Thurman, 1985). However, the percentages of HPO EfOM extracted from our effluent samples are similar to previous studies with treated municipal wastewaters that showed the HPO fraction to account for 26e40% of the EfOM (Baker, 2001; Krasner et al., 2009; Park et al., 2010; Pernet-coudrier et al., 2008). An additional 9e12% of the EfOM in our samples was extracted in the transphilic (TPI) fraction (Table 2). The TPI fraction is typically not quantified during natural DOM characterization. Our distributions of EfOM in the TPI fractions were also consistent with previous observations that have shown TPI fractions in wastewater EfOM to be as high as 20% (Krasner et al., 2009; Pernet-coudrier et al., 2008). The majority (>50%) of our EfOM was not retained by either the HPO or the TPI resin, but recovered in the hydrophilic (HPI) fraction (Table 2). Thus, one primary difference between DOM in wastewater effluent and natural systems is the much smaller proportion of organic matter that is operationally defined as hydrophobic. Our separation of EfOM fractions highlights a shortcoming of conventional resin extraction techniques for full characterization of organic matter from effluent sources. Resin extraction techniques that have been so effective for isolating natural DOM only capture between 30 and 50% of the organic matter in effluent samples. We focused our EfOM characterization efforts on the HPO and TPI fractions so that we could undertake comparisons of EfOM characteristics with characteristics of natural DOM that had been isolated with the same technique. HPO constitutes the bulk of organic matter in natural systems and it has been well characterized because of its potential role in phototransformation reactions, photooxidant generation and contaminant transport (Dalrymple et al., 2010; Guerard et al., 2009; Haitzer et al., 2003). Our resin isolation results (Table 2) do highlight the large fraction of hydrophilic organic matter that is present in wastewater effluent and that ultimately may be extracted using preconcentration techniques (Pernet-coudrier et al., 2008). We note that cation exchange resin treatment of our organic matter extracts reduced the iron and aluminum concentrations to less than 2.6 mg/mgC. We anticipate that these amounts of bound metals will not impart a shift in the size or optical characteristics of our organic matter, based on their similarity to the background levels of iron and aluminum in the work of Pullin et al. (2007).
3.2.
Organic matter size
Size exclusion chromatography showed isolated EfOM to have distinctly different characteristics from natural DOM. First, the chromatograms for the wastewater samples were multimodal, with two characteristic features (Fig. 1): (i) the presence of 5 (HPO) or 6 (TPI) distinct ‘peaks’ between 100 and 1000 Da, and (ii) the presence of a separated peak at 50,000 Da in CT2 and CT4. In contrast, size exclusion chromatograms of natural DOM are typically unimodal, with a single broad peak between 100 and 10,000 Da (Cabaniss et al., 2000; Chin et al., 1994).
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Fig. 1 e Size exclusion chromatograms of hydrophobic (left panels) and transphilic (right panels) fractions of CT2 (upper panels) and CT3 (lower panels) EfOM. Note that the chromatograms have different y-axis scales to show details of the peak distributions. Peak numbers shown on graph, hydrophobic fractions did not show peak 4.
Second, the number-averaged (Mn) and weight-averaged (Mw) molecular weights were smaller than observed for natural DOM. Our number- averaged molecular weights ranged from 290 to 360 Da for the HPO fraction (Table 3) whereas hydrophobic natural DOM typically has Mn values between 713 and 1360 Da (Cabaniss et al., 2000). Likewise, the Mw values for HPO EfOM ranged from 570 to 670 Da (Table 3), lower than the 1000e2310 reported for NOM (Cabaniss et al., 2000). Our measured polydispersities were between 1.5 and 2.2 (Table 3) which are consistent with reported data for natural sources of DOM, despite our overall lower averaged molecular weights. There was little variation between the average sizes of organic matter in our HPO and TPI fractions with the exception that weight-averaged molecular weights were about 100 Da higher on average for the HPO fraction than for the TPI fraction. Our EfOM size characteristics were similar to previous studies of EfOM molecular weights that reported Mn and Mw molecular weights of 400e500 and 650e850, respectively (Imai et al., 2002; Park et al., 2010). Thus, the HPO and TPI fractions of EfOM have lower averaged molecular weights than the corresponding fractions of organic matter isolated from natural sources. We excluded the 50,000 Da peak from our calculations of molecular weights for the CT2 and CT4 HPO fractions reported in Table 3. Isolated high molecular weight peaks have been observed in previous size exclusion chromatograms for
wastewater organic matter and are thought to consist of polysaccharide materials (Her et al., 2002; Imai et al., 2002). Inclusion of the high molecular weight peak in our molecular weight calculations greatly skews the average values, yielding a molecular weight (Mw) value of 4160 Da and polydispersity of 13.51 for CT2 and respective values of 1530 Da and 5.25 for CT4. The multimodal distribution of EfOM sizes (Fig. 1) followed a similar pattern for all wastewater samples. Such multimodal
Table 3 e Size characteristics of effluent organic matter in the 50e3200 Da range. Mw (Da)
Polydispersity
Hydrophobic fraction CT1 302.9 CT2 293.5 CT3 359.1 CT4 287.2 CT5 346.2
664.9 566.9 641.4 586.6 669.8
2.2 2 1.79 2.04 1.93
Transphilic fraction CT1 304.5 CT2 300.5 CT3 391.9 CT4 300.0 CT5 374.1
501.2 493.7 578.9 458.4 557.3
1.65 1.64 1.48 1.53 1.49
Location
Mn (Da)
290
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distributions of organic matter size have been observed in previous work investigating wastewater organic matter characteristics (Her et al., 2002; Imai et al., 2002). In our study, the size exclusion chromatograms consistently showed five peaks in the HPO fractions for all samples and six peaks in the TPI fractions (Fig. 1) between 1000 and 100 Da. Certain chromatograms did not have a sharp peak at all locations, but did exhibit a shoulder, indicating that a small peak was present. We calculated the molecular weight at the maximum intensity of each peak to examine whether any trends in peak masses were observed between effluent locations (Table S1, Supporting Information). The molecular weights of each of the chromatogram peaks were all very similar (5 to 20 Da) across the different treatment plants within and between organic matter fraction types. The appearance of discrete peaks in the low molecular weight organic matter suggests the presence of distinct organic byproducts/biomolecules in the wastewater treatment process. Furthermore, the fact that the peak molecular weights were remarkably consistent among the samples, despite the variation in treatment plant size and technologies employed (Table 1), suggests common characteristics of organic matter in treated wastewater effluent sources. We examined absorbance spectra of each peak of the size exclusion chromatograms to identify whether they may suggest similarities in chemical characteristics of EfOM between treatment plants. We collected absorbance scans from 220 to 350 nm whenever a peak was detected in a chromatogram (peak numbers are shown in Fig. 1). Peak 6 was not detected by the instrument, so absorption spectra were not available. The shape of the absorbance curves for peaks 1 through 4 in both the HPO and TPI fractions were very similar (peak numbers shown on Fig. 1). Each showed a steep, exponential-like decrease in absorbance with increasing wavelength from 220 nm and with a small shoulder at about 272 nm (Fig. S1, Supporting Information). Peak 5 differed from peaks 1e4 in the absence of the 272 nm shoulder and a steeper monotonic decrease in absorbance from 220 nm to 350 nm. To quantify decreases in peak absorbance spectra, the absorbance at 220 nm was divided by the absorbance at 272 nm (Table S2, Supporting Information). Typical values for peaks 1 through 4 ranged from 1.9 to 3.4, whereas the ratio for peak 5 ranged between 1.7 and 60. A sharper slope in absorbance for peak 5 is consistent with the organic matter in peak 5 being characterized by less carbonecarbon bond conjugation (little absorbance at longer wavelengths). This interpretation is reasonable given that the organic matter in peak 5 had a molecular weight of only 170 Da.
3.3.
Organic matter fluorescence
The EEMS of wastewater effluent fractions showed fluorescence signatures that are not found in natural DOM extracts. The EEMS for all samples and fractions of EfOM showed fluorescence in regions A and C that are characteristic of natural DOM; however, EfOM EEMS also showed fluorescence in regions B and T1 (Fig. 2). Comparisons of our EEMS with natural DOM samples are hampered somewhat by the bounds of conventional analyses that truncate excitation wavelengths below 250 nm. Nevertheless, there is some evidence
that peaks T1 and B are not typically present in natural DOM since studies of organic matter without 250-nm wavelength truncation showed low fluorescence intensity in the regions of peaks B and T1 (Baker, 2001; Holbrook et al., 2005). Suwannee River shows very little fluorescence in the Peaks B and T1 regions (Chen et al., 2003). To our knowledge, all published EEM data of microbial end member DOM is truncated at excitation wavelengths above 250 nm, so we were not able to undertake comparisons with these natural DOM sources. We observed different characteristics between the EEMS of the HPO fractions and TPI fractions for our samples (Fig. 2). Within our samples, the EEMS spectra of all HPO fractions showed similar characteristics to each other, and likewise the TPI fractions showed similar characteristics to each other. One primary difference between the HPO and TPI fractions was the occurrences of peaks A and T. In the HPO fraction, EEMS peaks A and T1 did not show distinct separation. This was particularly evident for the HPO fraction from CT5 (Fig. 2) in which fluorescence intensity was high through the range of excitation wavelengths between 230 and 250 nm and emission wavelengths between 330 and 450 nm. In contrast to the HPO fraction, the EEMS for the TPI fractions had a reduced fluorescence in the T1 region, and a distinct peak in the A region. The qualitative differences in the EEMS matrix contours were substantiated quantitatively with the integration of the fluorescence peak regions (Table 4). The HPO fractions had a smaller percentage contribution of peak C, as compared to the TPI fraction. Previous work has shown that peaks A and C are the two fluorescence regions for fulvic acids (Baker, 2001; Chen et al., 2003; Hudson et al., 2007). Since peaks A and C are both larger for the TPI EEMS, two explanations are possible: either fulvic acids are more strongly retained by the XAD4 resin, or fulvic acids are retained equally by both resins, and other fluorophores are more strongly retained on the DAX8 resin. We were unable to compare our regional fluorescence distributions to prior studies of EfOM because our choices of wavelength boundary differed from prior studies (Chen et al., 2003) as a result of excitation wavelength truncation at 230 nm to eliminate excess noise. Fluorescence index measurements indicated that the extracted EfOM was similar in characteristics to microbially derived natural DOM. All fluorescence indices fall within the ranges of the natural organic matter continuum (Table 5). In addition, all fluorescence indices were above 1.5, indicating that the organic matter from the all five treatment plants are closer in characteristics to NOM at the microbial end of the continuum (fluorescence index, 1.9; McKnight et al., 2001), regardless of hydrophobicity. The indices of the TPI fractions of organic matter are larger than those of the HPO fraction, thus indicating that the TPI fraction had fluorescence index characteristics somewhat closer to natural DOM microbial source end members than did the HPO fraction.
3.4.
Optical analyses
Like the fluorescence characterization, SUVA values also indicated that the EfOM had similar aromaticity to microbially derived natural organic matter of microbial origin. The SUVA values for all of the organic matter fractions were between 1.5 and 3 L mg1 m1, with the exception of the CT2 TPI fraction
w a t e r r e s e a r c h 4 6 ( 2 0 1 2 ) 2 8 4 e2 9 4
291
Fig. 2 e EEMS of CT2 (upper panels) and CT5 (lower panels) treatment plants. Hydrophobic fractions are shown in the left panels and transphilic fractions are shown in the right panels. The letters on the CT2 hydrophobic EEM represent fluorescence regions (Coble, 1996).
(Table 5). SUVA values of natural DOM typically range from 1.7 to 5.2 L mg1 m1 with lower values reported for organic matter from microbial sources (Weishaar et al., 2003). The SUVA values for our extracted EfOM were consistent with SUVA measurements from prior EfOM characterization studies (Imai et al., 2002; Krasner et al., 2009; Park et al., 2010). We observed the SUVA value of the HPO fractions for each treatment plant to be larger than their corresponding TPI fraction. Such differences suggest that the HPO fractions may have more aromatic character than the TPI fractions.
Table 4 e Percentage distribution of different EEMs regions by the FRI method. Location
B (%)
T1 (%)
A (%)
T2 (%)
C (%)
Hydrophobic fraction CT1 10.31 CT2 10.81 CT3 9.46 CT4 7.86 CT5 7.85
36.12 45.31 39.85 34.4 40
37.03 27.27 34.2 37.88 35.49
5.49 6.07 5.42 6.04 5.11
11.06 10.54 11.07 13.81 11.56
Transphilic fraction CT1 10.41 CT2 11.86 CT3 10.72 CT4 8.52 CT5 9.55
21.18 31.85 24.87 26.47 27.53
47.2 33.46 41.36 44.5 41.11
5.57 7.8 6.43 4.83 6.58
15.64 15.29 16.62 15.68 15.24
While the SUVA values for individual EfOM samples indicated similarity to natural organic matter from microbial origins, the pooled data suggest some differences in optical characteristics between organic matter from effluent sources and natural sources. Prior work has indicated a strong correlation between natural DOM size (Mw) and molar absorptivity at 280 nm with a slope of 3.5 Da cm M1 (Aldrich humic acid omitted) (Chin et al., 1994). Our pooled effluent organic matter samples also showed a correlation between Mw and molar absorptivity at 280 nm, but exhibited a much lower slope of 0.9 Da cm M1 (Fig. S2). Such a lower slope suggests that EfOM has fewer chromophore groups per ‘molecule’ than does natural organic matter. The absorbance spectral slopes of our isolated EfOM, as quantified with the E2/E3 ratio, also were similar to natural DOM from microbial sources. Our EfOM E2/E3 ratios (Table 5) fell with the range of values between 3.1 and 7 that have been reported for natural DOM (Dalrymple et al., 2010). Unlike the other optical characteristics, E2/E3 ratios do not exhibit a clear separation between microbial and terrestrial end member sources; however, low E2/E3 ratios are consistent with organic matter structures, such as lignin precursors, that can form charge transfer complexes that are manifest in the long wavelength ‘tail’ of organic matter absorption spectra (Del Vecchio and Blough, 2004). Thus, the high E2/E3 ratios that we observed for our isolated EfOM samples would suggest the lack of terrestrial organic matter precursors and hence, characteristics that were more similar to natural organic matter of microbial origin.
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Table 5 e Fluorescence and optical properties, and metal contents of effluent organic matter. Location
pH
Fluorescence index
SUVA (L mg1 m1)
E2/E3
Fe (mg mg1 C )
Al (mg mg1 C )
Hydrophobic fraction CT1 CT2 CT3 CT4 CT5
3.93 2.43 3.18 1.9 2.57
1.57 1.69 1.57 1.71 1.64
2.09 2.08 3.04 2.09 2.81
5.44 4.85 5.33 5.69 4.98
NM 0.97 1.4 NM 0.64
NM 0.26 0.65 NM 0.22
Transphilic fraction CT1 CT2 CT3 CT4 CT5
2.92 2.55 3.11 2.54 2.57
1.83 1.94 1.7 1.93 1.82
2.06 0.86 1.58 1.71 2.33
5.73 4.05 4.79 5.28 4.82
NM 2.6 0.37 NM 0.45
NM 1.6 0.28 NM 1.7
3.5.
Treatment plant intercomparison
Overall, our characterization of isolated EfOM indicated great similarity between treatment plants. A few differences were observed. The CT1 and CT4 extracted effluent had a greater fraction of EfOM that was hydrophobic in character (Table 2); however, the physical and chemical characteristics of these EfOM fractions fell within the range of the other three plants. CT2 and CT4 did show the presence of a high molecular weight organic matter (50,000 Da) peak in the HPO fraction which may contribute to the greater fraction of organic matter mass in the HPO fraction for these samples. The EEMS fluorescence regional integration of CT2 showed a much higher T:A ratio in the HPO fraction and higher T:A ratio in the TPI fraction that the other plants (Table 4). Such a difference was not observed for CT4, suggesting that the differing fluorescence characteristics for CT2 were not a result of the apparent high molecular weight component. Fluorescence indices were somewhat lower for CT1 and CT3 in the HPO fraction than the other plants (Table 5). SUVA values were greater for the HPO fractions of the CT3 and CT5 plants (Table 5). These results suggest that there is a somewhat greater range in characteristics for the hydrophobic than the hydrophilic fractions of effluent organic matter. The small differences in EfOM characteristics among our samples could not be linked generally to plant operating conditions. CT2 and CT4 both utilize mechanical aerators in activated sludge treatment, whereas the other plants deliver oxygen via bubble diffusers. It is possible that differences in shear characteristics between the two aeration treatments could cause the release of extracellular polymeric substances from floc in the presence of mechanical aerators. High SUVA values have been associated with advanced nitrogen removal, however, we did not observe a clear trend in our dataset. CT3 and CT4 both employ advanced nitrogen removal, yet only CT3 showed high SUVA in the hydrophobic fraction. CT5 also had a high SUVA value for the HPO fraction, although advanced nitrogen removal is not an explicit treatment step at that plant. Finally, some change in EfOM characteristics might have been expected with ultraviolet disinfection at the CT3 treatment plant. Studies of drinking water disinfection suggest that UV treatment reduces the SUVA value of finished waters and may reduce the proportion of lower molecular
weight organic matter (Lamsal et al., 2011). No obvious trend of lowered SUVA values (Table 5) or higher molecular weights (Table 3) were observed for the CT3 plant, relative to the other treatment plants. Nor, did the use of primary clarification (or not) appear to indicate trends in any EfOM characteristics. Thus, wastewater treatment plant operating conditions appear to have only a minor effect on EfOM characteristics.
3.6.
Environmental significance
The results of our comprehensive characterization of extracted EfOM from a number of treatment plant sources highlight closer similarities in EfOM characteristics to natural DOM from microbial sources than from terrestrial sources. Consequently, by most of our organic matter characterization techniques, EfOM appears to group closer to the microbial end member of the natural organic matter continuum. The one exception is the occurrence of A and T peaks in EfOM that are apparently not observed in DOM from natural sources. That most EfOM characteristics are similar to microbial DOM sources, indicates that streams receiving significant (e.g., >10% total flow) effluent discharges will contain DOM with a mixture of characteristics since DOM in stream systems is typically dominated by allochthonous, terrestrial sources of organic matter. What implications that this may have with respect to fate of chemical contaminants released in the effluent is yet unknown, as little work has been done to evaluate how sorption coefficients or indirect photolysis rates may be similar to, or different from, natural sources of DOM. Certainly, the similarity between EfOM characteristics from treatment plants that are representative of those found nationwide in the US suggests that understanding of EfOM effects on stream ecological and contaminant fate processes may be readily translated to other systems with wastewater inputs.
4.
Conclusions
Our results show that effluent organic matter characteristics differ from terrestrial organic matter that is typically present in small to moderately sized rivers.
w a t e r r e s e a r c h 4 6 ( 2 0 1 2 ) 2 8 4 e2 9 4
1. Wastewater effluent discharges introduce organic matter with distinctly different characteristics than typically present in small river systems. First, the hydrophobic and transphilic components of EfOM are closer in characteristics to natural DOM derived from microbial sources; whereas, natural stream water DOM tends to be dominated by terrestrial sources. 2. Second, wastewater EfOM contains between 60 and 80% material that is operationally defined as hydrophilic organic matter for which the environmental implications (e.g. carbon cycling, contaminant fates) are unknown. 3. Our results show that extracted effluent organic matter samples from plants have very similar characteristics. This work, which establishes beginning ranges of EfOM characteristics, highlights some questions about the ranges in EfOM characteristics that remain unanswered. For example, more study is required to evaluate the effects of factors such as influent type (e.g., large industrial contributions), treatment plant residence time, and stormwater contributions may have on the characteristics of EfOM.
Acknowledgment Funding for this project was provided by Connecticut Sea Grant project number PD-10-07 and by the UConn Environmental Engineering Program and Civil and Environmental Engineering Department. We thank the following wastewater treatment plant operators for permission and assistance in collecting samples that were used in this work: Arnie Bevins, Gene Ely, Kenneth Pelzar, Marylee Santoro, and Tom Tyler. We also thank two anonymous reviewers for their thoughtful comments and critique of an earlier version of this manuscript.
Appendix A. Supporting information Supporting information associated with this article contains: (i) organic matter sizes for SEC chromatogram peaks, (ii) absorption spectra at peaks of size exclusion chromatograms, (iii) absorbance intensity ratios for SEC chromatogram peaks, and (iv) parameters for fluorescence regional integration method, and (v) molar absorptivity at 280 nm plotted against weight-averaged molecular weight and can be found in the online version at doi:10.1016/j.watres.2011.10.010.
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Available online at www.sciencedirect.com
journal homepage: www.elsevier.com/locate/watres
Colloid retention at the meniscus-wall contact line in an open microchannel Yuniati Zevi a, Bin Gao b, Wei Zhang c, Vero´nica L. Morales a, M. Ekrem Cakmak d, Evelyn A. Medrano e, Wenjing Sang a, Tammo S. Steenhuis a,* a
Department of Biological & Environmental Engineering, Riley-Robb Hall, Cornell University, Ithaca, NY 14853-5701, USA Department of Agricultural and Biological Engineering, University of Florida, Gainesville, FL 32611-0570, USA c National Exposure Research Laboratory, Ecosystems Research Division, U.S. Environmental Protection Agency, Athens, GA 30605, USA d Department of Environmental Engineering, Cukurova University, Adana 01330, Turkey e Water Resources Engineering, Technical University of Delft, Delft, The Netherlands b
article info
abstract
Article history:
Colloid retention mechanisms in partially saturated porous media are currently being
Received 18 February 2011
researched with an array of visualization techniques. These visualization techniques have
Received in revised form
refined our understanding of colloid movement and retention at the pore scale beyond what
15 September 2011
can be obtained from breakthrough experiments. One of the remaining questions is what
Accepted 23 September 2011
mechanisms are responsible for colloid immobilization at the triple point where air, water,
Available online 25 October 2011
and soil grain meet. The objective of this study was to investigate how colloids are transported to the air-water-solid (AWS) contact line in an open triangular microchannel, and
Keywords:
then retained as a function of meniscus contact angle with the wall and solution ionic
Contact angle
strength. Colloid flow path, meniscus shape and meniscus-wall contact angle, and colloid
Retention
retention at the AWS contact line were visualized and quantified with a confocal micro-
Capillary forces
scope. Experimental results demonstrated that colloid retention at the AWS contact line was
Air-water-solid interface
significant when the meniscus-wall contact angle was less than 16 , but was minimal for the
Contact line
meniscus-wall contact angles exceeding 20 . Tracking of individual colloids and computa-
Colloids
tional hydrodynamic simulation both revealed that for small contact angles (e.g., 12.5 ),
Microchannel
counter flow and flow vortices formed near the AWS contact line, but not for large contact angles (e.g., 28 ). This counter flow helped deliver the colloids to the wall surface just below the contact line. In accordance with DLVO and hydrodynamic torque calculations, colloid movement may be stopped when the colloid reached the secondary minimum at the wall near the contact line. However, contradictory to the prediction of the torque analysis, colloid retention at the AWS contact line decreased with increasing ionic strength for contact angles of 10e20 , indicating that the air-water interface was involved through both counter flow and capillary force. We hypothesized that capillary force pushed the colloid through the primary energy barrier to the primary minimum to become immobilized, when small fluctuations in water level stretched the meniscus over the colloid. For large meniscus-wall contact angles counter flow was not observed, resulting in less colloid retention, because a smaller number of colloids were transported to the contact line. ª 2011 Elsevier Ltd. All rights reserved.
* Corresponding author. Tel.: þ1 607 255 2489. E-mail addresses:
[email protected] (Y. Zevi),
[email protected] (B. Gao),
[email protected] (W. Zhang),
[email protected] (V.L. Morales),
[email protected] (M. Ekrem Cakmak),
[email protected] (W. Sang),
[email protected] (T.S. Steenhuis). 0043-1354/$ e see front matter ª 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2011.09.046
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1.
w a t e r r e s e a r c h 4 6 ( 2 0 1 2 ) 2 9 5 e3 0 6
Introduction
Colloids facilitate the transport of a wide range of contaminants through the vadose zone and into groundwater. Some typical contaminants that experience colloid-facilitated transport and are of great concern for human and ecological health include: radionuclides, pesticides and trace metals (Saiers and Hornberger, 1999; Williams et al., 2006). While most colloid transport studies have been carried out under saturated conditions, fewer studies have characterized colloid retention processes in unsaturated soils. These studies found that, in addition to retention mechanisms occurring in saturated soils, e.g., pore straining, and attachment to grain surfaces by the overall Derjaguin-Landau-Verwey-Overbeek (DLVO) energy minima and/or surface chemical heterogeneity or roughness that produce locally attractive sites on an overall repulsive surface (Johnson et al., 2010), the presence of a gaseous phase in unsaturated soils forms other potential retention sites such as the air-water interface (AWI) and the air-water-solid (AWS) contact line (Bradford and Torkzaban, 2008). Early work in unsaturated processes reported colloid retention at the airwater interface (Wan and Wilson, 1994) and in thin water films enveloping grains (Wan and Tokunaga, 1997; Veerapaneni et al., 2000). Research in the last decade has utilized microscopy advances and identified two new retention sites: immobile water zones (Gao et al., 2006) and the AWS contact line, which is defined as the triple point where air, water and grain approach each other. Colloid attachment at the AWS contact line, which is the focus of this study, has been observed in soil media (Crist et al., 2004, 2005; Zevi et al., 2005, 2009) and in surrogate of soil pores (e.g., microchannels) (Lazouskaya et al., 2006; Lazouskaya and Jin, 2008). Capillary and DLVO forces have both been considered to explain colloid retention at this interface by Crist et al. (2005), Gao et al. (2008), and Shang et al. (2008). This study sets out to explain why colloids adhere at the AWS contact line. To date several studies have indicated that a variety of colloids attach at the AWS contact line, especially when the DLVO forces do not favor colloid attachment at the solid-water interface (SWI) (Scha¨fer et al., 1998a; Sirivithayapakorn and Keller, 2003; Auset and Keller, 2006; Lazouskaya et al., 2006; Crist et al., 2004, 2005; Zevi et al., 2005, 2009). The studies by Crist et al. (2004, 2005) were likely the first time that the colloids were visualized to attach at the AWS contact line. Other earlier studies might have also observed this phenomenon, but attributed it to the attachment at the air-water interface (Wan and Wilson, 1994; Sirivithayapakorn and Keller, 2003). Although a controversy exists regarding the role of evaporation at the AWS contact line (Steenhuis et al., 2005; Wan and Tokunaga, 2005), since then the colloid attachment at the AWS contact line has been observed by Lazouskaya et al. (2006), Lazouskaya and Jin (2008), Gao et al. (2008), Morales et al. (2009), and Zevi et al. (2009). Despite these observations there has not been a satisfactory explanation on how colloids can approach the often energetically repulsive solid-water and/or air-water interfaces. Hydrodynamic forces have been speculated to be responsible for supplying the necessary external energy to overcome the repulsive energy exerted at the AWS contact line (Lazouskaya et al., 2006). However, according to Shang et al.
(2008), drag forces at realistic groundwater flow rates are too weak to thrust suspended colloids through the energy barrier at the AWS contact line so that capillary force can more permanently immobilize the colloids. Although the source of energy that colloids use to approach this energetically unfavorable retention site remains uncertain, experimental observations indicate that colloid immobilization at the AWS contact line ubiquitously occurs. For now, its mechanistic understanding remains elusive. The studies of Gao et al. (2008) and Shang et al. (2008) indicate that capillary forces are generally at least two orders of magnitude greater than the system’s DLVO forces. A detailed theoretical description of DLVO and capillary forces is presented in the auxiliary material. O’Brien and van den Brule (1991) suggested that colloid attachment on a solid substrate is controlled by capillary force of the film covering the solid, and can be determined by the liquid-colloid and liquid-substrate contact angles. Despite this critical observation for colloid retention in unsaturated systems, many studies have explained colloid attachment by employing the DLVO theory exclusively (e.g., Auset and Keller, 2006; Crist et al., 2004, 2005; Lazouskaya et al., 2006; Scha¨fer et al., 1998a,b; Sirivithayapakorn and Keller, 2003). Numerical models for colloid retention in unsaturated porous media mainly consider the air-water interface as an additional retention site as compared to saturated media (Corapcioglu and Choi, 1996; Massoudieh and Ginn, 2007; Chen, 2008). Specific model for colloid attachment at the AWS contact line are unavailable; therefore, additional work in this area is needed. Shi et al. (2010) simulated a moving airwater interface in a microchannel. Other studies have used the finite element analysis software COMSOL in colloid transport studies (Cakmak et al., 2008; Kemps and Bhattacharjee, 2009; Torkzaban et al., 2007, 2008; Bradford and Torkzaban, 2008). The objective of this study was to experimentally visualize and quantify how colloids flowing through a simplified “soil pore” (i.e., microchannel) are retained at the AWS contact line under a wide range of DLVO forces (varied by solution ionic strengths) and capillary force (varied by the air-water interfacial shapes exhibiting different meniscus-wall contact angles).
2.
Materials and methods
2.1.
Colloid attachment experiment in a microchannel
The colloids used were hydrophilic (water contact angle ¼ 12 ), negatively charged, 1-mm yellow fluorescent synthetic polystyrene microspheres (Polysciences, Inc., Warrington, PA). Colloid suspensions of 9.1 106 microspheres/ mL were prepared in the solution of 0.001% Rhodamine B dye at three ionic strengths (i.e., 1 mM, 50 mM, and 100 mM NaCl), respectively. The very low concentration of Rhodamine B dye was added to allow for imaging the water phase and does not influence colloid behavior (Zevi et al., 2006, 2009). The colloidfree solutions of identical solution chemistry with that of the colloid suspensions were used as background solutions. Prior to each experiment the colloid suspensions were sonicated in
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an ultrasonic bath (Bransonic B-52, Branson Ultrasonics, Danbury, CT) for 5 min to ensure monodispersivity. The principal components of the experimental setup consisted of an open triangular microchannel, syringe pumps, and a confocal microscope, as illustrated in Fig. 1a. The microchannel made of borosilicate glass consisted of an isosceles triangular shape carved canal with a 90-degree angle, a length of 54-mm and a wall height of 5-mm (Fig. 1b and c). The channel piece was mounted on the confocal microscope stage with an inclination of 2 to prevent ponding of inflow liquid. The channel inlet and outlet were composed of small capillary reservoirs (10-mm long) that dispersed the flow from the tube to the channel cross-sectional area (Fig. 1b). Push and pull syringe pumps were used (Orion Sage Pump model 341B, Thermo Scientific; KDS model 120, KD Scientific, Inc., Holliston, MA) to maintain a constant and unpulsed flow into and out of the channel. The visualization system was a Leica TCS SP2 laser scanning confocal microscope equipped with an HC PL FLUOTAR CS objective (5.0 magnification) with a numerical aperture of 0.15, yielding a resolving power of 0.97 mm/pixel (Fig. 1a). Before commencing the experiment, the microchannel was cleaned by sonication in the ultrasonic bath with a low concentration of anionic glassware cleaner (Alconox Inc.) in distilled water for 15 min, rinsed with distilled water for five times followed by rinsing with ethanol and distilled water again, and finally dried at 65 C for 30 min. The cleaned microchannel mounted under the microscope was initially empty, and then the testing liquid was added and removed by the pumps, as illustrated in Fig. 1a. Two inlet syringes were
prepared containing the background solution and the colloid suspension, respectively. The microchannel was initially filled with the background solution using the inlet syringe pump at a high flow rate (10 mL/h) until the water front reached the outlet reservoir, whereupon the outflow pump was turned on to begin withdrawing the liquid until the water level reached equilibrium. Once the water level reached a uniform height of 2 mm along the channel, the flow rate of both pumps were reduced to 1 mL/h in order to further lower the water height to approximately 1.5 mm. Typically 15e30 min were required to obtain a continuous and uniform flow. The average linear flow velocity under a steady-state flow condition was 0.12 mm/s. Once at the steady-state, a pulse of the colloid suspension was introduced into the uninterrupted inflow by rapidly exchanging the inlet tubing from the syringe containing the colloid-free background solution to the second inlet syringe containing the colloid suspension. Initial experiments demonstrated that the contact angle of the meniscus (q) in the channel could be increased by covering the channel with a 24 50 mm glass microscope slide (Fisher Scientific). Therefore, for selected experiments, the contact angle at the AWS contact line was altered by covering the channel. For such experiments, the glass cover was placed on the top of the microchannel after the flow became uniform and continuous, and the rest of experimental procedures were the same as that described above. The length of the microchannel covered by the glass slide was 50 mm, representing 93% of the total coverable channel surface area. Experiments were conducted in duplicate or triplicate for conditions of open (without a cover) or closed channel (with a cover) at solution ionic strengths of 1, 50 and 100 mM NaCl (see Table 1 for a detailed experimental list). The parameter combinations resulted in six different treatments to compare a broad range of DLVO and capillary forces. To study the effect of covering the channel on the contact angle without flowing water, one static treatment was added where the channel was covered and the flow was stopped after the pulse of colloids was injected. With the confocal microscopy system, simultaneous imaging of the colloids, Rhodamine B dyed water and the microchannel was achieved using excitation/emission spectral channels set at 488/700 nm (Argon laser), 543/650 nm (HeNe laser), and transmitted light, respectively. A more detailed description of the instrumental setting is given by Zevi et al. (2005). During each experiment, two types of image sequences were recorded at a distance of 3 cm away from the inlet. Initially, a “xyz” image stack was recorded to construct a 3-D
Table 1 e Summary of experiments performed. Experiment
Fig. 1 e (a) Schematic diagram of the experimental setup, (b) Detailed view of the microchannel, (c) Schematic cross section of the microchannel.
Series 1 Series 2 Series 3 Series 4 Series 5 Series 6 Series 7
Solution
Condition
1 mM NaCl 50 mM NaCl 100 mM NaCl 1 mM NaCl 50 mM NaCl 100 mM NaCl 1 mM NaCl
Open channel Open channel Open channel Closed channel Closed channel Closed channel Open and closed channel
Flow Number of rate replicates 1 ml/h 1 ml/h 1 ml/h 1 ml/h 1 ml/h 1 ml/h Static
3 2 3 2 2 2 2
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cross section of the channel at the beginning of the experiment and subsequently used to measure the meniscus contact angle at the microchannel walls. From these images the meniscus-wall contact angle was determined using 2D tools, as described in detail by Zevi et al. (2009). The measurements were repeated for at least four different cross sections because of observed variation in the profile within a single channel. Additional measurements of the contact angle were made at succeeding stages of each experiment: prior to the injection of the colloid pulse, during the experiment if the meniscus shape was observed to have changed, and at the end of the experiment. During the experiment, but after the meniscus was well-established, a series of time sequenced images (xyt series) were collected to monitor the accumulation of colloids at the AWS contact line, with a frequency of 1e2 images per second. From these images, colloid retention was quantified by analyzing consecutive images in a given region using ImageJ software, as carried out by Zevi et al. (2005, 2009). In addition, two series of time sequenced images (xyt series) from two experiments described earlier (one with large contact angle of 28 at solution ionic strength 100 mM NaCl and one with a small contact angle of 12.5 at solution ionic strength 1 mM NaCl) were sufficiently clear to allow for tracing the path of the colloids near the AWS contact line with manual tracking by the ImageJ software. The period of particle tracking was 4 min for the experiment with the 12.5-degree contact angle and 40 min for the experiment with 28-degree contact angle. With the right calibration value for xy and interval time, the xy coordinates, travel distance and velocity of one colloidal particle between two successive images were automatic recorded by simply clicking the colloidal particle on the image window. Multiple colloids were tracked for each experiment. Based on these data, an overlay of the original stack and traced trajectories were generated, with a different color applied for each traced colloid.
2.2. Computational simulation of flow field in the microchannel The three-dimensional steady-state flow field in the triangular channel with 54 mm of length was simulated by numerically solving the Stokes equations using COMSOL Multiphysics v3.5a software package (COMSOL, Inc., Burlington, MA). Vp ¼ mV2 v
(1)
V$v ¼ 0
(2)
where p is the fluid pressure (Pa), n is the velocity of the water (m/s), m is the dynamic viscosity of the water (0.001 Pa s). The upper curved boundary, which accounts for the airwater interface (i.e., the water meniscus), was set as a slip boundary condition (i.e., the normal component of velocity is zero and the tangential component of total stress is zero, Zhang et al., 2010). The channel walls were specified as a noslip boundary condition (i.e., zero velocity). The contact angle between the meniscus and the channel wall (q) was set at 12.5 and 28 . The flow was induced by specifying a fluid velocity of 0.12 mm/s at the inlet and zero pressure at the outlet of the channel.
2.3. Calculations of DLVO and hydrodynamic forces and torques Colloid retention is greatly dependent on the forces and torques that act on the colloids (Bradford and Torkzaban, 2008; Bradford et al., 2011). Thus, DLVO forces (FDLVO) were calculated for colloids interacting with the SWI and AWI. The calculations of hydrodynamic drag force (FD), applied hydrodynamic torques (Tapplied), and DLVO adhesive torques (Tadhesive) were performed for the colloids attached on the SWI in a typical Poiseuille type flow field, and the detailed equations are provided in the auxiliary material. The balance of Tapplied and Tadhesive determines whether the colloids could remain immobilized on the SWI.
3.
Results
The results from a total of 14 experimental runs (6 treatments with 2 or 3 replicates, experimental series 1 to 6, Table 1) were used to analyze the shape of the meniscus by way of the meniscus-wall contact angle, and the degree of colloid attachment at the AWS contact line. The system variables included the solution ionic strength (1, 50, and 100 mM) and the meniscus-wall contact angle, which was controlled partly by the presence and absence of a channel cover. By employing this set of combinations, the DLVO forces were altered by the solution’s ionic strength and the capillary force by the meniscus-wall contact angle.
3.1.
Meniscus and contact angle effects
Examples of reconstructed cross section from the “xyz” image stack of the microchannel and fluid, depicting the shape of the meniscus 30 min after the colloid pulse was initiated, are given for the contact angle less than 10 (Fig. 2a), between 10 and 20 (Fig. 2b), between 20 and 40 (Fig. 2c), and greater than 40 (Fig. 2d). The red color in these figures was the Rhodamine B dyed water. Green yellowish colloids near the contact line in Fig. 2a and b were elongated, which is typical for bright objects in the reconstruction of cross-sectional image stacks. The yellow spots especially notable in the center of Fig. 2a were due to the artifacts of the confocal microscope. In Fig. 2eeh, the longitudinal channel view of colloid attachment near the AWS contact line is shown at the same 30-minute time period. Colloids can be found as green yellowish dots either in a straight line at the AWS contact line or scattered in the “red” water. The ones at the contact line were stationary, while the others were in motion. The experiment with a meniscus-wall contact angles less than 10 (Fig. 2e) had the greatest amount of colloids retained at the contact line. For the contact angle between 10 and 20 colloids were retained to a less extent (Fig. 2f), but for larger contact angles colloid attachment at the AWS contact line was almost negligible (Fig. 2g and h). Colloids were not deposited at any other locations in the microchannel.
3.2. Colloid retention at the AWS contact line as a function of contact angle and ionic strength Despite the careful experimentation and cleaning procedures, contact angles between the wall and the meniscus varied
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Fig. 2 e Examples of the shape of the meniscus and colloid retention at the AWS contact line 30 min after the colloid pulse was introduced for different meniscus-wall contact angles: (a, e) < 10 , (b, f) 10e20 , (c, g) 20e40 , (d, h) > 40 .
greatly even for the same treatments. In general the covered experiments had greater contact angles (ranging from 42 to 63 , Fig. 3aec) than uncovered experiments with contact angles between 7 and 29 (Fig. 3def). We cannot readily explain these differences in contact angles, but it is assumed to be a result of slight differences in the channels used, cleanliness and heterogeneity of the surface, and the hysteresis of the contact angle (Johnson and Dettre, 1964; Lam et al., 2002; Johnson et al., 2010). Even after the flow of water was stopped, the contact angle remained at 30 in the covered microchannel (Series 7, Table 1). We should also note that after running the experiments for several hours the contact angles decreased and eventually approached the static value of around 10 . Colloid retention (Fig. 3) at 30 min after the injection of colloid pulse in all experiments with the meniscus-wall contact angles over 20 was less than 0.3 mm2/mm (defined as the area occupied by colloids per length of the contact line).
These observations are independent whether or not the microchannel was covered. For two uncovered experiments at 1 mM and 100 mM ionic strengths where the contact angle was over 20 , the colloid retention was small (under 0.3 mm2/ mm), similar to all covered experiments. For contact angles less than 16 colloid retention was significant and up to 3.0 mm2/mm for the 1 mM experiments (Fig. 3d). The retention of colloids at the injection time of 30 min is re-plotted as a function of the average contact angle (Fig. 4a) and ionic strength (Fig. 4b). Fig. 4a indicates that for the meniscus-wall contact angles smaller than 29 the colloid retention decreased with increasing contact angles. Colloid retention at a contact angle greater than 20 was minimal and less than 0.3 mm2/mm as noted above. In addition, there was a linear decrease in colloid retention with increasing ionic strength for the subgroup of 10e20 (R2 ¼ 0.96) for which sufficient data is available. Zevi et al. (2009) observed a similar trend for colloid retention at the AWS contact line in partially saturated sand
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Fig. 3 e Retention of the colloids at the AWS contact line for experiments with a closed (aec) and open channel (def) in the 1, 50 and 100 mM NaCl solutions. The average contact angle, q, plus or minus the standard deviation is presented for each microchannel experiment.
pack, which was caused by increased attachment at the solidwater interface with increasing ionic strength. Since there was no colloid attachment at the microchannel wall, the above explanation of Zevi et al. (2009) is not applicable here. We will discuss possible causes later.
3.3. Flow field effects on colloid retention at the AWS contact line The flow paths of colloids approaching the AWS contact line were investigated experimentally and computationally. The experimentally observed flow patterns are given in Fig. 5a for the meniscus-wall contact angle of 28 and in Fig. 5b for the 12.5-degree contact angle. The movies with the real time pathways are included in the auxiliary materials (Section B). When the contact angle was 12.5 10 colloids approached the channel wall during a 4-minute observation time (Fig. 5b). Once the colloids arrived in the vicinity of the wall they moved in the opposite direction to the general flow field before becoming immobilized at the AWS contact line. In contrast, for the 28-degree contact angle only three colloids approached
the wall during an observation period of 40 min (Fig. 5a). Unlike the case with the smaller meniscus-wall contact angle (Fig. 5b) there was no counter flow, and the colloids settled at the wall while moving forward. The computational results are given at Fig. 6a and b for the same set of the meniscus-wall contact angles (i.e., 12.5 and 28 ). The normalized velocity field was illustrated by the red arrows. The boundaries of the channel were made visible by a gray color. The simulated flow for the contact angle of 12.5 showed the counter flow close to the AWS contact line, (Fig. 6a) whereas no counter flow was observed in the simulation for the contact angle of 28 (Fig. 6b).
3.4.
Calculated forces and torques acting on colloids
The calculated DLVO forces for colloid-SWI interactions indicated that with increasing ionic strength the magnitude of the attractive forces at the primary minimum ðFDLVO 1min Þ decreased, whereas the attractive forces at the secondary minimum ðFDLVO 2min Þ increased (Table 2 and Figure A2). There is a formidable repulsive force for the colloids interacting with the AWI
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301
maximum as discussed later. As the colloids approach the AWS contact line (i.e., decreasing H ), the ratio of decreases, which increases the colloid retenTapplied =Tadhesive 2min tion at the SWI (Table 3). For example, at H ¼ 50 mm, the colloids will not remain stationary on the SWI, but roll away at all ionic strengths because Tapplied is greater than Tadhesive 2min (Table 3), which explains the absence of colloid retention at the channel wall far away from the AWS contact line. The torque analysis also predicts that the colloid retention at the secondary minimum will increase with increasing ionic strength (Table 2 and Table 3). Since this prediction does not agree with the decreased colloid retention with increasing ionic strength at the AWS contact line for the meniscus-wall contact angles of 10e20 as discussed above (Fig. 4b), some of the assumptions should be adjusted. This discrepancy indicates that at the AWS contact line the secondary minimum is not the sole colloid retention mechanisms. More discussion is provided in a later section.
4.
Fig. 4 e Colloid retention at 30 min after colloid addition: (a) plotted as a function of average contact angle; (b) plotted as a function of ionic strength.
(Figure A2). The applied hydrodynamic forces (FD) and torques (Tapplied) for a Poiseuille type flow field decreases as the colloid approaches the AWS contact line (i.e., smaller depths of the AWS wedged pore space [H], Table 3), suggesting that the colloid will be more likely retained due to reduced hydrodynamic drag. The calculated Tapplied values in Table 3 are comparable to the adhesive torques at the secondary minimum ðTadhesive 2min Þ, but three to six orders of magnitude smaller than the adhesive torques at the primary minimum ðTadhesive 1min Þ (Table 2). However, colloids could not be retained at the primary minimum due to the repulsive force at the primary maximum (Table 2) unless other force (e.g., capillary force) is involved to push the colloid through the primary
Fig. 5 e Trajectories of colloidal particles for (a) 28-degree contact angle and (b) 12.5-degree contact angle.
Discussion
The effectiveness of capturing colloids at the AWS contact line decreased with increasing meniscus-wall contact angles (Fig. 4a). Analogous to colloid deposition to a collector under saturated conditions, the capture of colloids at the AWS contact line depends on how colloids approach the interface (collision efficiency) and how effective these colloids are retained by the interface (attachment efficiency) (Yao et al., 1971; Brown and Abramson, 2006; Cakmak et al., 2008). In the following sections, based on visual evidence, computational simulations, and theoretical considerations, we discussed first the transport of colloids to, and subsequently their retention at the AWS contact line. Colloid transport to the AWS contact line was enhanced by counter flow along the walls as shown by the computational simulations (Fig. 6) and experimental path tracking (Fig. 5 and the movies in the auxiliary material section B). Counter flow occurred for small contact angles but not for larger contact angles (Figs. 5a and b, 6a and b). These observations are in agreement with the experiments of Lazouskaya et al. (2006) and Lazouskaya and Jin (2008) and with the latticeBoltzmann simulations of Shi et al. (2010) for a moving contact line in a microchannel with a relatively small contact angle. Although their experiments were transient due to a moving wetting front and different with our steady flow, in both cases colloids reach the interface as a result of counter flow that exists in these microchannels as can be seen in Movie-3 of Lazouskaya and Jin (2008) and Movie 1 and Movie 2 in the auxiliary material section B. Supplementary video related to this article can be found at doi:10.1016/j.watres.2011.09.046. Additionally, the counter flow (e.g., backward flow, and flow vortices) was also observed in the COMSOL simulations (Torkzaban et al., 2008) at the grainegrain contacts and the AWS contact line (Zhang et al., 2010), and numerical simulation at the moving contact line in a capillary tube (Sheng and Zhou, 1992). The lattice-Boltzmann saturated flow simulations by Li et al. (2010) in pore structures derived from real
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Fig. 6 e Computational simulations of flow field in a 54-mm long microchannel: (a) 12.5-degree meniscus-wall contact angle and (b) 28-degree meniscus-wall contact angle. (i) The cross section of microchannel with water phase (in pink color) (ii) The flow field along the channel (iii) The enlargement of one flow field (red arrows represent flow direction). (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)
porous media indicates that counter flow not only occurs at the AWS contact line in our simplified model soil pore, but also at the grainegrain contacts in the realistic soil pore structures. Of interest is that Wan and Tokunaga (2005) attributed the observed differences of colloid retention between a closed and open porous media to the effect of evaporation. Based on the experiments that investigated the evaporation effect between covered and uncovered experiments, Steenhuis et al. (2005) found that the effect of evaporation was negligible. In addition, using the theory presented by Morris (2001) we can easily prove that evaporation cannot
account for the observed difference in contact angles between the closed and open channels in this study. Once the colloid is transported to the wall, it can only stay stationary when the hydrodynamic drag forces are less than a combination of other forces (e.g., DLVO, capillary and friction forces). As explained above, colloid retention at the AWS contact line is not solely determined by the DLVO forces, because an increased colloid retention with increasing ionic strength is not observed as predicted by the theory (Fig. 4b). Instead, there was a slight decrease of colloid retention with increasing ionic strength for the contact angle of 10e20 . The
Table 2 e DLVO forces (FDLVO) normalized with the colloid radius (r [ 0.5 mm) and adhesive torques (Tadhesive) for colloids interacting with the solid-water interface (SWI) (see Appendix for detailed calculations).a IS (mM)
Primary minimum FDLVO 1min =r
1 50 100
(mN/m)
11.6 1.67 1.30
x (nm) 0.35 0.35 0.35
Tadhesive 1min
Primary barrier (N m)
8.26 1017 6.23 1018 4.46 1018
FDLVO max =r
(mN/m)
1.56 3.85 3.80
Secondary minimum
x (nm) 5 1 1
FDLVO 2min =r
(mN/m)
6.48 105 7.71 103 1.79 102
x (nm)
Tadhesive (N m) 2min
123 13 8
5.16 1024 3.02 1021 9.28 1021
DLVO a FDLVO 1min ¼ attractive DLVO force at the primary energy minimum, Fmax ¼ repulsive DLVO force at the primary energy maximum, DLVO F2min ¼ attractive DLVO force at the secondary energy minimum, x ¼ separation distance, Tadhesive ¼ DLVO adhesive torque at the primary 1min energy minimum, and Tadhesive ¼ DLVO adhesive torque at the secondary energy minimum. 2min
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Table 3 e Applied hydrodynamic torque (Tapplied), hydrodynamic drag force (FD) normalized with the colloid radius (r), and adhesive the ratio of Tapplied =T2min for 1-mm colloid in different depths of the air-water-solid wedged pore space (the flow depth [H] ranging from 1 mm to 50 mm) (see Appendix for detailed calculations).a q ¼ 12:5
H (mm)
FD =r (mN/m)
q ¼ 28
v ¼ 1:19 103 m=s;
v ¼ 9:17 104 m=s;
DP=Lc ¼ 87:2 Pa=m
DP=Lc ¼ 48:0 Pa=m
Tapplied (N m)
1 mM 1 5 10 20 50
6.99 6.29 1.33 2.73 6.92
7
10 106 105 105 105
2.45 2.20 4.65 9.54 2.42
22
10 021 1021 1021 1020
FD =r (mN/m)
Tapplied Tadhesive 2min
47.4 427 901 1849 4695
50 mM 0.08 0.73 1.54 3.16 8.02
Tapplied (N m)
100 mM 0.03 0.24 0.50 1.03 2.61
3.85 3.46 7.31 1.50 3.81
7
10 106 106 105 105
1.35 1.21 2.56 5.25 1.33
22
10 1021 1021 1021 1020
Tapplied Tadhesive 2min 1 mM
50 mM
100 mM
26.1 235 496 1018 2580
0.04 0.40 0.85 1.74 4.42
0.01 0.13 0.28 0.57 1.44
a q ¼ water contact angle of the meniscus, v ¼ average water velocity in the channel, and DP/Lc ¼ pressure gradient.
meniscus-wall contact angle is a more important factor that determines the colloid retention at the AWS contact line, which for small contact angles produces a counter flow slowing down the colloids so that the hydrodynamic torque is less than what would be expected from the Poiseuille type of laminar flow. As observed in the movies, at the large distance (e.g., H > 20 mm) from the AWS contact line and large contact angle, the attractive DLVO forces at the secondary minimum are not sufficiently strong to permanently keep the colloids at the surface (Table 3). An exception existed for the 100 mM ionic strength featured in the Movie 2 where a few colloids that approached the glass wall were retained. This observation agrees with the three-dimensional particle tracking coupled with the lattice-Boltzmann flow simulations in porous media by Li et al. (2010), which showed that colloids may intermittently stop moving by a weak association with the grain surface via the secondary minimum. Applied hydrodynamic drag forces are not great enough to push the colloid through the primary energy barrier (Tables 2 and 3), and certainly not large enough to push against the capillary force (Figure A3 in Auxiliary material section A.4). Since colloids for the small contact angles were attached firmly on the wall after the experiments were completed, indicating that they had a strong bond with the glass surface that is the characteristics of the primary minimum. Colloids can reach the primary minimum by either a force on the colloid that is greater than the repulsive force of the primary energy barrier, or by reducing the energy barrier force. Since enhanced colloid retention due to a decrease in the primary barrier repulsive force has been reported for the cases of nanoscale surface chemical heterogeneity or roughness, we cannot rule out this attachment mechanism (Duffadar and Davis, 2007; Duffadar et al., 2009; Johnson et al., 2010). However, because the same glass surface was used in all the experiments, a more likely candidate for pushing the colloids through the existing energy barrier would be the capillary force associated with the meniscus. However, the capillary force only acts when the meniscus is “stretched” over the colloid as in Figure A1 (Auxiliary material section A.2). Since the drag forces are not sufficient, we hypothesized that small movements of the meniscus due to slight changes in flow are the main factor responsible for stretching the meniscus over
and thus initiating the action of the capillary force on the colloid. Since the triple point where the meniscus attach to the surface is reluctant to move (i.e., contact angle hysteresis; Johnson and Dettre, 1964; Lam et al., 2002), any decrease in flow will decrease the contact angle, but more importantly, it will force the meniscus to stretch over the colloid as shown in Figure A1. At this point capillary force will start acting. Capillary force can be decomposed into two force components, i.e., the lateral force that pushes the colloid back in solution and the normal force that pushes the colloid against the wall (see Auxiliary material Section A.4). The normal capillary force will push the colloid through the primary barrier ranging from 1.6 mN/m for the 1 mM ionic strength to 3.8 mN/m for the 100 mM ionic strength (Table 2). Even for small protrusions (i.e., the colloid protrudes 10% of its diameter through the meniscus) the normal capillary force is around 45 mN/m for contact angles ranging from 12 to 45 (Figure A3 in Auxiliary material section A.4). Once the colloid is pushed through the energy barrier it will reach the primary minimum at 0.35 nm from the wall. For the colloid to stay at the primary minimum, the lateral capillary force has to be smaller than the friction force of the colloid with the glass wall and the attractive force at the primary minimum. The friction force is equal to the product of the friction coefficient and the sum of the normal components of the capillary force and the DLVO force at 0.35 nm (Table 2, Figure A3). By setting the friction force equal to the lateral force (Auxiliary material section A.5 Eq. (A12)), we can calculate the minimum friction coefficient that is needed for the colloid to just stay at the primary minimum. The minimum friction coefficients for contact angles of 12.5 and 28 and the three ionic strengths are shown in Fig. 7. For the 12.5-degree contact angle the friction coefficient (between the colloid and glass) must be at least of 0.06 for the colloid to remain at the wall near the AWS contact line for an ionic strength of 1 mM and a protrusion distance through the meniscus (d ) at 3% of the colloid radius r (i.e., d/r ¼ 0.03) (Fig. 7). For the colloid to remain at the wall at greater ionic strengths, the friction coefficient has to be at least 0.22 when the colloid protrudes farther through the meniscus (Fig. 7). The Engineering Toolbox (2011) states that the friction coefficient of lubricated glass/ metal is between 0.2 and 0.3. In other words, it is plausible that
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that of Lazouskaya et al. (2006). They observed as well that the shape of meniscus is related to the amount of colloid retention with larger contact angles having less retention.
5.
Fig. 7 e Minimum friction coefficient required for colloid retention at the AWS contact line dependent on meniscuswall contact angles (q [ 12.5 and 28 ) and ionic strength (1 mM, 50 mM, and 100 mM NaCl solutions) at the primary minimum (0.35 nm). For the x-axis, r is the colloid radius, and d is the distance that the colloid protrudes out of the meniscus.
the friction coefficient for wet glass and microspheres falls between 0.2 and 0.3, indicating that the colloids would be retained at the AWS contact line when the contact angle is 12.5 . This is in agreement with the experimental data in Figs. 3 and 4a where colloids remain attached at the wall for the meniscuswall contact angles smaller than 16 . Moreover, since at low ionic strength, the minimum friction coefficients required for attachment are less than for greater ionic strength, we can expect the greatest attachment efficiency at smaller ionic strength. This is collaborated by experimental data in Fig. 4b. At the meniscus-wall contact angle of 28 , for colloids to remain at the wall the required minimum friction coefficient depicted in Fig. 7 are generally above 0.35 for ionic strength of 50 and 100 mM. Only at the 1 mM ionic strength, a lower required minimum friction coefficient makes it possible for colloids to be retained when they barely protrudes through the meniscus. Although the exact value of static glass/colloid friction coefficient is unknown, for the 28-degree contact angle the required minimum friction coefficients is much greater than for the smaller contact angles, and likely more than what can be expected for colloids with wet glass (Fig. 7). Therefore, the attachment efficiency could be close to zero. The data in Fig. 4 underpin these theoretical findings with independently measured experimental results. Additionally, although the above analysis was performed for a hydrophilic colloid, capillary force is also important for the retention of hydrophobic colloids since capillary force overwhelms any other surface forces. For increased hydrophobicity of the colloid (i.e., a greater colloid-water contact angle), the colloid may be required to protrude farther out of the meniscus so that the capillary force can pin the colloid onto the solid surface since the direction and magnitude of capillary force are dependent on the colloid-water contact angle and the protrusion distance (d ) (Auxiliary material section A.4 Eq. (A9) and (A11)). For extremely hydrophobic colloids, they may become more easily associated with the airwater interface. Finally, our results are also in agreement with
Conclusions
Retention of colloids at the AWS contact line was observed experimentally, which was attributed to the presence of the contact angles of water meniscus and solid surface less than 16 . The small meniscus-wall contact angle induced counter flow near the triple point where water, wall and air meets (i.e., the AWS contact line). For larger contact angles counter flow was not experimentally observed. Computational simulations were in agreement with the experimentally observed flow pattern. Once colloids were in the close proximity of the wall near the AWS contact line, colloids were retained in the secondary minimum when hydrodynamic drag was small. It was hypothesized that small changes in average flow rate decreased the contact angle, thus stretching the meniscus over the colloids located in the secondary minimum. Capillary forces associated with the stretched meniscus could push colloids through the energy barrier into the primary minimum.
Acknowledgments This research was supported by funding from the USDANational Research Initiative (project 2005-03929 and 200835102-04462), Binational Agricultural Research and Development Fund, Project No. IS-3962-07, and the National Science Foundation (project 2006-0635954). The authors acknowledge the expert guidance of Carol Bayles, the manager of the Cornell University Biotechnology Center’s Microscopy and Imaging Facility.
Appendix. Supplementary material Supplementary material associated with this article can be found, in the online version, at doi:10.1016/j.watres.2011.09.046.
references
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Cation exchange during subsurface iron removal D. van Halem a,*, D.H. Moed a, J.Q.J.C. Verberk a, G.L. Amy a,b,1, J.C. van Dijk a a b
Delft University of Technology, Faculty of Civil Engineering and Geosciences, Stevinweg 1, 2628 CN Delft, The Netherlands UNESCO-IHE, Westvest 7, 2611 AX Delft, The Netherlands
article info
abstract
Article history:
Subsurface iron removal (SIR), or in-situ iron removal, is an established treatment tech-
Received 20 April 2011
nology to remove soluble iron (Fe2þ) from groundwater. Besides the adsorptive-catalytic
Received in revised form
oxidation theory, it has also been proposed that the injection of O2-rich water onsets the
11 October 2011
exchange of adsorbed Fe2þ with other cations, such as Ca2þ and Naþ. In sand column
Accepted 12 October 2011
experiments with synthetic and natural groundwater it was found that cation exchange
Available online 15 November 2011
(NaþeFe2þ) occurs during the injection-abstraction cycles of subsurface iron removal. The Fe2þ exchange increased at higher Naþ concentration in the injection water, but
Keywords:
decreased in the presence of other cations in the groundwater. Field results with injection
Cation exchange
of elevated O2 concentrations (0.55 mM) showed increased Fe removal efficacy; the oper-
Drinking water treatment
ational parameter V/Vi (abstraction volume with [Fe]0. In the experiments, [Fe2þ] is allowed to breakthrough, in order to calculate the dimensionless retardation factor R. RFe þ 1 is calculated from a Vi corresponding to the pore volume when the tracer (electrical conductivity) is C/C0 ¼ 0.5, and V is the number of pore volumes that can be abstracted with iron concentrations below (C/C0)Fe ¼ 0.5. In the case of a column study, the oxidation zone is limited to the size of the column, i.e., 1 pore volume, making the calculation of V/Vi redundant:
Table 1 e Summary of conducted experiments for synthetic and natural waters: The injection water quality and the groundwater water quality. Experiment
Location
Injection water
Groundwater
A B C D E A B C D E
Laboratory
0.28 mM O2, 5 mM NaHCO3 0.01 M NaHCO3, No O2 0.1 and 0.5 M NaCl, No O2 No Na, No O2 No Na, 0.28 mM O2 0.28 mM O2, 5 mM NaHCO3 5 mM NaHCO3, No O2 0.1 M NaCl, No O2 No Na, No O2 No Na, 0.28 mM O2
pH ¼ 6.9 (0.02), T ¼ 20 C (0.1), 0.1 mM Fe2þ (0.01), 5 mM NaHCO3
WTP
pH ¼ 7.2 (0.01), T ¼ 11 C, 0.1 mM Fe (0,01), 3.3 uMMn, 1.07 mMCa, 0.52 mM Na, 0.28 mM Si and 0.14 mM SO4
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311
Fig. 4 e Fe breakthrough in duplicate columns after injection of water containing 0.28 mM O2 and 5 mM NaHCO3 for columns loaded with synthetic and natural groundwater (C0 [ 0.1 mM Fe2D).
RWell
V Vi FeC=C ðPVÞFeC=C 0 0 ¼ 10RColumn ¼ 1 V ðPVÞTracerC=C 0 Vi TracerC=C
(4)
0
The Fe retardation during abstraction for the regular injection-abstraction cycle is shown for both the synthetic and natural groundwater columns in Fig. 4. The tracer passed the columns at PV ¼ 1, but elevated Fe concentrations were not observed until 30 and 8 pore volumes for the synthetic and natural columns, respectively. The Fe retardation in the natural groundwater columns (RFe ¼ 9) was measured to be lower than for synthetic water (RFe ¼ 42). The reduction in removal efficiency can be explained by the presence of competing cations in the natural groundwater, binding to the same sites as Fe2þ. For instance, Ca2þ is known to inhibit the Fe2þ adsorption onto virgin and iron hydroxide coated sand (Sharma, 2001). When considering the synthetic groundwater column results it was calculated from the retardation factor (RFe ¼ 42) that the Fe retention was approximately 0.6 mmol/cycle. This value correlates well to the 0.6 mmol Fe calculated from the O2 consumption in the columns. Based on this finding it may be concluded that all O2 consumption was used for Fe2þ oxidation, however, there is still the question of whether cation exchange plays a (catalyzing) role. Adsorbed Fe2þ could either directly (heterogeneously) oxidize on the surface, like proposed with the catalytic adsorptive-oxidation mechanism, or it could exchange with other cations before oxidizing to Fe3þ hydroxides. During a regular injection-abstraction cycle, the Fe:O2 mass balance does not differentiate between these two mechanisms, as Fe2þ will always oxidize in the presence of O2 and be retained in the columns.
3.2.
Fe2þeNaþ exchange
After reaching steady state with the synthetic groundwater the columns were loaded for an injection phase without O2, but with elevated Naþ concentrations. Naþ was added as
Fig. 5 e Fe measurements during (A) injection of 0.01 M NaHCO3, Na-free or 0.5 M NaCl without O2 and (B) abstraction of synthetic groundwater (C0 [ 0.1 mM Fe2D), dotted lines represent results from PHREEQC (v2.15).
0.01 M NaHCO3 or 0.5 M NaCl. Additionally a control cycle was tested without the addition of any Naþ. It is noteworthy that the pH during injection with water containing low buffer capacities (absence of HCO 3 ) remained above pH 7.5. Fig. 5 shows the Fe concentrations that were measured during the injection phases. The dotted lines represent the results from the geochemical surface complexation model PHREEQC (v2.15; Parkhurst and Appelo, 1999). Fe desorption was measured from the column material at concentrations 1.5e25 times the C0 of 0.1 mM. The PHREEQC model results even show a higher desorbed concentration, indicating that during sampling the high peak concentration was missed. The amount of exchanged Fe was higher in the case of 0.5 M Naþ, resulting in subsequent higher removal efficiencies during abstraction (Fig. 5). RFe was measured to be 11 and 30, for 0.01 M and 0.5 M Naþ, respectively. The same experiment was conducted with 0.1 M NaCl, showing the same retardation as for 0.5 M NaCl (not shown, confirmed with PHREEQC), which means that maximum Naþ regeneration has occurred at 0.1 M NaCl. The retardation of Fe during this particular cycle provides information on the exchangeable Fe2þ capacity of the sand, as retention of approximate 0.4 mM Fe can be recalculated to a Cation Exchange Capacity (CEC) of 1.55 meq.kg1 (0.78 mmol Fe.kg1) for the column sand. This CEC value is lower than for
312
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average aquifer sand (10 meq.kg1; Appelo and Postma, 2005), as can be expected based on the absence of clay, organic carbon and iron oxides on this clean (silica) filter sand. The CEC value correlates well with the Fe2þ adsorption capacity of virgin sand of 0.74 mmol Fe.kg1 observed by Sharma et al. (1999). In that study the adsorption of Fe2þ onto clean sand (D ¼ 0.7e1.25 mm) at pH 7.0 in the absence of oxygen was investigated. When using the measured exchange value of 1.55 meq.kg1 in PHREEQC, the measured results were simulated well. It should be noted, though, that there was slightly more tailing of the Fe2þ breakthrough curve during abstraction in the measurements than the model, potentially due to kinetics and/or stagnant zones. These results show that also in the absence of O2, but in the presence of abundant Naþ, a retardation of Fe can be achieved to a certain level. The desorption of Fe during injection shows the exchange of attached Fe2þ with soluble Naþ, and subsequent retardation of Fe points towards the vise versa exchange of Fe2þ and Naþ. The occurrence of Fe2þeNaþ exchange, even on clean filter sand, can potentially play an important role during subsurface iron removal. However, this depends strongly on cation concentrations in the injection water and on the CEC of the aquifer material. The control cycle with Naþ-free water confirmed that the Fe2þ exchange was very low during injection (Fig. 5). Fe was subsequently not significantly delayed in the columns, with an RFe of 1.
3.3.
CIEX in natural groundwater
The role of CIEX in the multi-component environment of natural groundwater was studied in experiments using groundwater from a water treatment plant (WTP Loosdrecht) instead of synthetic groundwater. Under these natural conditions the injection cycles in the presence and absence of O2 and/or Naþ were investigated. Fig. 6A depicts the Fe desorption during the injection in the absence of oxygen. A peak concentration almost reaching the initial concentration of 0.1 mM was detected. At 0.1 M NaCl the Fe leaching from the columns was higher than for 5 mM Naþ, resulting in a slightly better Fe retardation during the abstraction phase (Fig. 6B). The Fe desorption after injection in the absence of Naþ was most likely caused by the cations in the natural groundwater present in the columns just before injection. It is noteworthy that the leaching in these natural groundwater columns was up to 20 times lower than in the synthetic groundwater columns, as can be seen from Fig. 5, not exceeding the background Fe concentration of 0.1 mM. The Fe retardation factors that were measured in all experiments for both synthetic and natural groundwater are summarized in Fig. 7, showing the decrease in efficiency in the natural groundwater columns compared to the synthetic groundwater columns. The RFe after injection with or without oxygen show a reduced value in the multi-component environment of natural groundwater (injection modes A, B, C, E). The Fe2þ exchange/adsorption is clearly limited by the presence of other cations, such as Ca2þ, resulting in retardation values below 12. In the absence of oxygen, the RFe was somewhat similar for both 5 mM and 0.1 M Naþ, i.e., CIEX between Naþ and Fe2þ is not significantly enhanced at higher Naþ. Also in the absence of Naþ during injection, the Fe2þ retardation was limited in natural groundwater compared to synthetic water
Fig. 6 e Fe measurements during (A) injection of 5 mM NaHCO3, 0.1 M NaCl or Na-free water without O2 and (B) abstraction with natural groundwater.
(injection mode E), as the RFe dropped from 33 to 8 in the multicomponent environment of the natural groundwater columns. It should be noted that although RFe between injection modes A, C and E correlate well, the Fe breakthrough trend looks different (not shown). After injection of Na-free water, the Fe
Fig. 7 e The measured Fe retardation factors for cycles with different injection modes with abstraction of either synthetic or natural groundwater.
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concentrations never reached below the detection limit, i.e., some Fe always passed the columns. This was observed in the synthetic and natural groundwater columns, indicating that some CIEX is needed to reach ultra low Fe concentrations at the beginning of an abstraction phase. In Fig. 8 the measurements for Na and Ca are depicted, showing some desorption of Ca during injection (Fig. 8A). For the duration of 9 pore volumes, the Na concentration increased during injection from the concentration in the groundwater (C/C0 ¼ 0.12) to the concentration in the injection water (C/C0 ¼ 1). The tailing of both curves indicates that CIEX proceeds between these cations. The abstraction period shows the reversed trend (Fig. 8B), as the Na concentration of the groundwater (C0) is reached after approximately 8 pore volumes. The Ca concentration is already restored after 6 pore volumes. It should be noted that Ca2þ complexation with natural organic matter (NOM) may have occurred as well, but because of the low binding contact time can be considered insignificant. It may be concluded that CIEX occurs, in the multi-component environment of natural groundwater, between a wide range of cations, resulting in retardation during the abstraction phase. The effect of CIEX is most predominant on Fe retardation, as Fe concentrations in the groundwater are relatively low compared to Ca2þ and Naþ.
3.4.
Fe2þ desorption theory
The column studies in the absence of O2 and/or Naþ confirmed that, in addition to the adsorptive-oxidation process (van Beek,
1985; Rott, 1985), CIEX occurs during subsurface iron removal. The Naþ concentrations of actual injection water will always be lower than the concentrations in the experiments. The experiments therefore overestimate the actual occurrence of CIEX during subsurface iron removal. However, the CEC of clean filter sand in the columns is much lower than that of actual aquifer sand (on average 10 times) resulting in lower CIEX. At the beginning of a full-scale injection cycle, the CIEX will happen in the presence of O2, resulting in immediate oxidation. While injection proceeds, the injection water will penetrate further into the aquifer, but the O2 front will lag behind, caused by the consumption of O2 during Fe2þ oxidation. In other words, the injected water front does not contain any O2, but contains cations for exchange with Fe2þ. The further the injection water flows into the aquifer, the larger the distance between the injected water front and the O2 front becomes. In this moving zone where O2 is absent, the process of CIEX prevails and Fe2þ desorbs from the soil material and travels deeper into the aquifer (Figs. 5A and 6A). In theory, these elevated Fe2þ concentrations never come in contact with the O2 in the injection water and will not oxidize in this cycle (Fig. 9). When abstraction starts and the flow direction is reversed, the desorbed Fe2þ passes the available adsorption sites on the soil grains closer to the production well and is removed through either adsorption or CIEX. In other words, the injection phase of subsurface iron removal mobilizes a part of the adsorbed Fe2þ through CIEX but does not subsequently oxidize the Fe2þ. One may state that this proportion of mobilized Fe2þ limits the system’s efficacy during the abstraction phase, as part of the adsorption sites will be occupied by the desorbed Fe2þ and not by the Fe2þ present in the groundwater. On the other hand, the exchange may also enhance SIR, as the desorbed Fe2þ is pushed deeper into the aquifer and concentrations may be lowered through the buffering capacity of the aquifer material. Whether the Fe2þ truly enhances or limits the efficacy of SIR is yet to be determined, because it all depends on the actual separation between the Fe2þ desorption front and the O2 front. The investigation of this front separation was not the focus of these column experiments and is recommended for future research.
3.5.
Fig. 8 e Na and Ca measurements during a normal injection (A) and abstraction (B) cycle with natural groundwater.
313
Injection elevated O2 concentrations
CIEX may play a role during subsurface iron removal, but it is the supply of O2 to the (im)mobile Fe2þ that determines the efficiency of the system. Appelo et al. (1999) concluded that increasing the oxidant concentration of the injected water would be useless as long as the efficiency is limited by the amount of exchangeable Fe2þ capable of consuming the oxidant. The observation that injected oxygen will not be used in the absence of available oxidizeable Fe2þ is valid, but there is also a reason why injection of higher O2 concentrations can increase the system’s efficacy. Namely, if the injection water contains higher O2 concentrations, the O2 front will not lag far behind the injected water front. The desorbed Fe2þ will then come in contact with O2 for oxidation and not leach out of the oxidation zone into the aquifer. Thus by injecting higher oxidant concentrations, the exchangeable Fe2þ fraction on the soil material can be utilized for oxidation and will contribute
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Fig. 9 e Schematic representation of the separation between O2 front and Fe2D desorption peak.
as hydrolyzed Fe3þ hydroxides with their exchangeable/ adsorptive surface area during the following abstraction phase. At water treatment plant Corle (Vitens Water Supply Company) there has been extensive experience with the injection of elevated oxygen concentrations into subsurface iron removal wells. In the past they injected 3000 m3 of drinking water containing 0.28 mM O2, but they have changed the operational mode to the injection of 2000 m3 with an O2 concentration of 0.55 mM. Although the O2 concentration increased with a factor of 1.9, the total O2 injection increased only with a factor 1.3, from 0.84 to 1.1 *103 mol. As a result of this operational change, the volumetric ratio for abstraction (V) and injection (Vi) has increased from an average 7 to 16. This immense and sudden efficiency increase was clearly caused by the operational change, as an increasing efficacy caused by successive cycles (Appelo et al., 1999) may not be expected at stabilized subsurface iron removal wells. WTP Corle calculates V/Vi based on the moment when Fe breakthrough starts, so the moment [Fe]>2 mM is registered.
Fig. 10 depicts the relation of the operational parameter V/ Vi for 12 subsurface iron removal wells after injection of 0.28 mM O2 and 0.55 mM O2. Although there was some variation between the results, considering these are operational data from 12 different wells, it can be concluded that on average the V/Vi increases by an approximate factor 2. This indicates that not the available (adsorbed and/or exchangeable) Fe2þ is limiting during injection, but the supply of O2 to the (in)soluble Fe2þ. The field measurements support the theory that increasing the oxidant concentration has a positive effect on the subsurface removal of Fe, but they do not prove that it is actually the Fe2þ desorption front that is targeted by the higher O2 dose. The field results point towards the assumption that Fe2þ is abundantly available on the soil material. In conclusion, CIEX may occur in the aquifer during injection, but does not seem to limit the system’s efficiency when injecting elevated oxygen concentrations.
4.
Conclusions
In sand column experiments with synthetic and natural groundwater it was found that cation exchange (Naþ-Fe2þ) occurs during the injection-abstraction cycles of subsurface iron removal. The Fe2þ exchange increased at higher Naþ concentration in the injection water, but decreased in the presence of other cations in the groundwater. Field results with injection of high O2 concentrations indicated that not the exchangeable Fe2þ on the soil material is the limiting factor during injection, but it is the supply of O2 to the available Fe2þ.
Acknowledgements
Fig. 10 e The operational V/Vi ratio of 12 wells at WTP Winterswijk with (0.55 mM) and without (0.28 mM) the injection of 47 molmL3 technical oxygen.
The research presented in this article was supported by the InnoWATOR grant (Innovation Program for Water Technology) of Agentschap NL. The authors would like to thank Erik van der Pol and Hans Bergevoet of Vitens Water Supply Company for their support.
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references
Appelo, C.A.J., Drijver, B., Hekkenberg, R., de Jonge, M., 1999. Modeling in situ iron removal from ground water. Ground Water 37 (6), 811e817. Appelo, C.A.J., Postma, D., 2005. Geochemistry, Groundwater and Pollution, second ed. Balkema, Rotterdam. Boochs, P.W., Barovic, G., 1981. Numerical-Model Describing groundwater treatment by Recharge of Oxygenated water. Water Resources Research 17 (1), 49e56. Braester, C., Martinell, R., 1988. The Vyretox and Nitredox methods of in situ treatment of groundwater. Water Science and Technology 20, 149e163. Grombach, P., 1985. Groundwater treatment in situ in the aquifer. Water Supply 3 (1), 13e18. Hallberg, R.O., Martinell, R., 1976. Vyredox - in situ purification of groundwater. Ground Water 14 (2), 88e93. Hiemstra, T., van Riemsdijk, W.H., 2007. Adsorption and surface oxidation of Fe(II) on metal (hydr)oxides. Geochimica et Cosmochimica Acta 71, 5913e5933. Mettler, S., Abdelmoula, M., Hoehn, E., et al., 2001. Characterization of iron and manganese precipitates from an in situ groundwater treatment plant. Ground Water 6, 921e930. Mettler, S. (2002) In-situ removal of iron from groundwater: Fe(II) oxygenation, and precipitation products in a calcareous aquifer. PhD dissertation, Swiss Federal Institute of Technology, Zurich. Parkhurst D.L. and C.A.J. Appelo (1999) User’s guide to phreeqc (version 2) e a computer program for speciation, batchreaction, one-dimentional transport, and inverse geochemical calculations. Water-Resources Inverstigation Report 99-4259, US Geological Survey.
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Rott, U., 1985. Physical, chemical and biological aspects of the removal of iron and manganese underground. Water Supply 3 (2), 143e150. Rott, U., Meyer, C., Friedle, M., 2002. Residue-free removal of arsenic, iron, mangenese and ammonia from groundwater. Water Science and Technology: Water Supply 2 (1), 17e24. Rott, R., Friedle, M., 1985. Physical, chemical and biological aspects of the removal of iron and manganese underground. Water Supply 3, 143e150. Sharma, S.K., Greetham, M.R., Schippers, J.C., 1999. Adsorption of iron(II) onto filter media. Journal of Water SRT-Aqua 48 (3), 84e91. Sharma S.K. (2001) Adsorptive iron removal from groundwater. PhD dissertation, Wageningen University. Sung, W., Morgan, J.J., 1980. Kinetics and products of ferrous iron oxygenation in aqueous systems. Environmental Science & Technology 14, 561e568. Tamura, H., Kawamura, S., Hagayama, M., 1980. Acceleration of the oxidation of Fe2þ ions by Fe(III)-oxyhhydroxides. Corrosion Science 20, 963e971. van Beek, C.G.E.M., 1985. Experiences with underground water treatment in the Netherlands. Water Supply 3 (2), 1e11. van Halem, D., Olivero, S., de Vet, W.W.J.M., Verberk, J.Q.J.C., Amy, G.L., van Dijk, J.C., 2010. Subsurface iron and arsenic removal for shallow tube well drinking water in rural Bangladesh. Water Research 44, 5761e5769. van Halem, D., de Vet, W., Verberk, J., Amy, G., van Dijk, H., 2011. Characterization of accumulated precipitates during subsurface iron removal. Applied Geochemistry 26, 116e124. Wolthoorn A. (2003) Subsurface aeration of anaerobic groundwater; iron colloid formation and the nitrification process. Ph.D. dissertation, Wageningen University.
w a t e r r e s e a r c h 4 6 ( 2 0 1 2 ) 3 1 6 e3 2 6
Available online at www.sciencedirect.com
journal homepage: www.elsevier.com/locate/watres
Effects of ozone and ozone/peroxide on trace organic contaminants and NDMA in drinking water and water reuse applications Aleksey N. Pisarenko a, Benjamin D. Stanford a,b, Dongxu Yan a,c,d, Daniel Gerrity a,e, Shane A. Snyder a,c,* a
Southern Nevada Water Authority, Applied Research & Development Center, PO Box 99954, Las Vegas, NV 89193, USA Hazen and Sawyer, P.C., Raleigh, NC 27607, USA c University of Arizona, Chemical and Environmental Engineering, Tucson, AZ, USA d Layne Christensen Company, 3804 E. Watkins Street, Phoenix, AZ 85034, USA e Trussell Technologies, Inc., 6540 Lusk Blvd., Suite C274, San Diego, CA 92121, USA b
article info
abstract
Article history:
An ozone and ozone/peroxide oxidation process was evaluated at pilot scale for trace
Received 27 June 2011
organic contaminant (TOrC) mitigation and NDMA formation in both drinking water and
Received in revised form
water reuse applications. A reverse osmosis (RO) pilot was also evaluated as part of the
13 October 2011
water reuse treatment train. Ozone/peroxide showed lower electrical energy per order of
Accepted 15 October 2011
removal (EEO) values for TOrCs in surface water treatment, but the addition of hydrogen
Available online 25 October 2011
peroxide increased EEO values during wastewater treatment. TOrC oxidation was corre-
Keywords:
predicting contaminant removal. A decrease in N-nitrosodimethylamine (NDMA) forma-
N-Nitrosodimethylamine (NDMA)
tion potential (after chloramination) was observed after treatment with ozone and ozone/
Trace organic contaminant (TOrC)
peroxide. However, during spiking experiments with surface water, ozone/peroxide ach-
Ozone
ieved limited destruction of NDMA, while in wastewaters net direct formation of NDMA of
lated to changes in UV254 absorbance and fluorescence offering a surrogate model for
Ozone/H2O2
6e33 ng/L was observed after either ozone or ozone/peroxide treatment. Once formed
Ozone/peroxide
during ozonation, NDMA passed through the subsequent RO membranes, which highlights
Ozone/hydrogen peroxide
the significance of the potential for direct NDMA formation during oxidation in reuse
Pharmaceuticals
applications.
Endocrine
disrupting
compounds
ª 2011 Elsevier Ltd. All rights reserved.
(EDCs) Advanced oxidation process (AOP)
1.
Introduction
As the global population is projected to reach nine billion by 2050, water reuse and desalination will become a critical
water resource for much of the world (Bereta and Miller, 2010). As such, ensuring adequate removal of trace organic contaminants (TOrCs) of health concern in reclaimed waters, including endocrine disrupting compounds (EDCs) and N-
* Corresponding author. University of Arizona, Chemical and Environmental Engineering, 1133 E. James E. Rogers Way, Tucson, AZ 85721-0011, USA. E-mail address:
[email protected] (S.A. Snyder). 0043-1354/$ e see front matter ª 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2011.10.021
w a t e r r e s e a r c h 4 6 ( 2 0 1 2 ) 3 1 6 e3 2 6
nitrosodimethylamine (NDMA), becomes increasingly significant. While most pharmaceuticals and personal care products (PPCPs) and EDCs are not believed to pose significant human health threats at concentrations currently found in water (EPHC et al., 2008; Snyder et al., 2010), their presence is still a concern for many individuals and is the foundation for much of the resistance to widespread implementation of potable water reuse. Thus, there exists a need to study TOrC removal through various multi-barrier water reuse treatment processes. Various oxidation processes such as ozone, ozone/ peroxide, UV/H2O2, and non-thermal plasma have been reported to be effective in destroying TOrCs (Benotti et al., 2009a; Gerrity et al., 2010, 2011; Rosario-Ortiz et al., 2010; Rosenfeldt and Linden, 2004; Snyder et al., 2007; Wert et al., 2009a), but these processes are not equally able to remove all compounds. For example, one of the advantages of UV/H2O2 is the ability to remove NDMA and NDMA precursors in water (Kruithof et al., 2007) via photolysis and hydroxyl radical (OH) oxidation. This may be especially important in reuse applications where the eventual chlorination and/or chloramination of the finished water may result in more NDMA formation (PehlivanogluMantas et al., 2006). In contrast, ozone and ozone/peroxide have been implicated in direct formation of NDMA during oxidation of dimethylamine (DMA) and may in fact exacerbate NDMA concerns (Andrzejewski et al., 2008). Furthermore, Schmidt and von Gunten reported direct NDMA formation due to oxidation of N,N-Dimethylsulfamide by ozone in natural waters (Schmidt and Brauch, 2008; von Gunten et al., 2010). Despite the direct formation of NDMA, ozonation is extremely effective in reducing formation potential, which is the formation of NDMA following chloramination. Assuming direct NDMA formation can be controlled, it may be worthwhile to ozonate water intended for reverse osmosis (RO), as this water will require chloramination to control biological fouling of the membranes. In addition to the reductions in NDMA formation potential, ozonation can also be used to reduce organic fouling on microfiltration, ultrafiltration, and RO membranes due to the transformation of organic matter (Park et al., 2007; Pisarenko et al., 2011; Stanford et al., 2011; You et al., 2007; Zhu et al., 2010). Since RO is such an energy intensive process, preozonation has the potential to yield sufficient energy and cost savings to warrant its inclusion in advanced treatment trains. Therefore, it is important to not only evaluate the direct benefits of contaminant oxidation but also the benefits realized after integration into the larger treatment train, notably reduced RO breakthrough of contaminants and improved quality of the RO concentrate. The primary objective of this study was to evaluate TOrC destruction in drinking water and wastewater using a pilotscale ozone and ozone/peroxide technology. The efficacy of both ozone and ozone/peroxide were monitored to evaluate NDMA destruction, direct NDMA formation, NDMA formation potential, and the destruction of other TOrCs (e.g., pharmaceuticals and personal care products (PPCPs)). In order to narrow the scope of this research, a subset of the numerous compounds detected in previous occurrence studies was selected for evaluation. The indicator compounds were selected based on their magnitude and frequency of occurrence in water and wastewater (Snyder et al., 2007), varying
317
physical/chemical characteristics and resulting susceptibility to treatment (Snyder et al., 2007; Ternes et al., 2002; Westerhoff et al., 2005) and ease of analytical methods (Trenholm et al., 2009). Since bromate formation has historically been the most significant concern related to the use of ozone in water treatment (Orlandini et al., 1997; Wert et al., 2007), bromate was also monitored during the study. With respect to the advanced treatment trains, the project team also evaluated the integration of this technology as a possible pretreatment to an RO membrane system to minimize transport of trace contaminants and consequent impacts on the RO permeate. In this way, this technology is broadly evaluated for its application in reuse applications.
2.
Materials and methods
2.1.
Pilot-scale equipment and tested waters
For the wastewater tests, a primary-treated wastewater from the Las Vegas Valley was treated using a pilot-scale HYDRAsub Membrane Bioreactor (MBR) system (Hydranautics, Oceanside, CA) with hollow-fiber vacuum-type polyvinylidene fluoride (PVDF) membranes with a nominal pore size of 0.40 mm. The general MBR operating parameters were consistent with those previously described (Stanford et al., 2011) and are provided in the supplementary information (Table SI-I). Briefly, the MBR was operated at a mixed liquor suspended solids (MLSS) concentration of approximately 8000 mg/L, a solids retention time (SRT) of 12 days, and a hydraulic retention time (HRT) of 4 h. For the target compounds that were not completely removed by biological treatment, contaminant concentrations were monitored before and after ozonation using the pilot-scale HiPOx system (APTwater, Pleasant Hill CA). The HiPOx system was used in this study because it recently received California Department of Public Health Title 22 certification for disinfection in wastewater and water reuse applications. In addition benchscale ozonation experiments were performed by spiking ozone into MBR filtrate or secondary effluent samples at similar ozone doses, following a previously reported method (Stanford et al., 2011). The pilot scale RO skid (Hydranautics, CA) was used to monitor breakthrough of organic contaminants through ESPA2-4040 RO membranes. The surface charge of a clean ESPA2 membrane is typically 20 to 40 mV with an isoelectric point at pH of 4.0. The pilot skid was operated at 52% recovery using a single array of 6 elements. The feed water was adjusted to pH 6.8 using 50% sulfuric acid (Brenntag, NV) and spiked with bulk sodium hypochlorite (obtained from the wastewater treatment plant) and food grade ammonium chloride, 99% (Brenntag, NV) to form residual monochloramine. For the drinking water tests, Colorado River water (CRW) from Lake Mead, NV was used as the source water for the pilot-scale HiPOx and bench-scale RO membrane testing. The bench-scale RO testing consisted of two identical GE Osmonics (Minnetonka, MN) Sepa-CF cross-flow membrane flat-sheet cell holders to process the raw and oxidized waters using a protocol previously described (Stanford et al., 2011). The Colorado River water received at the pilot plant was prechlorinated (0.8 mg/L of free chlorine) by the local water
318
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utility to control Quagga Mussel growth in the intake structures. The experimental water initially contained a free chlorine residual less than 0.2 mg/L, but the water was also recirculated for 24 h in a 3000-gallon batch tank to remove residual chlorine. The feed water was also filtered by a 5micron cartridge filter prior to filling the tank. A recirculation rate of 10 gpm allowed for complete turnover of the batch tank nearly 5 times in the 24-hour period. At the end of 24 h, a spike solution of the analytes dissolved in laboratory grade water was added to the batch tank, and the water was recirculated for an additional 24 h. The analytes were not dissolved in a solvent in order to prevent the introduction of potential OH scavengers. Control samples were collected at the beginning and end of the experiments to evaluate consistency in the target compound concentrations. The HiPOx pilot unit is capable of operating in a variety of modes and configurations, at flow rates of 10e25 gpm, and at ozone doses of up to 15 mg/L. The pilot was fed either using liquid oxygen feedstock or a high-purity (99.9%) oxygen gas to generate up to 10% ozone in dry gas. A concentrated hydrogen peroxide solution (34%) was purchased from (EnviroTech Chemical Services, Modesto, CA), diluted to a 1e2% working solution, and injected immediately before ozone injection in the HiPOx reactor. In addition to the injection ports and static mixers, the HiPOx pilot contains a 60-gallon pipeline contactor with numerous sampling ports that allow sample collection for hydraulic residence times ranging from 0 to 5.5 min at a flow rate of 10 gpm. The MBR filtrate and Colorado River water were exposed to the following ozone doses by adjusting oxygen flow, generator power, and monitoring percent ozone in the dry gas: 0.6, 1.5, 3.0, 6.0, and 10.0 mg/L. The transfer efficiency was typically high (>95%) so the applied ozone dose was approximately equal to the transferred ozone dose.
2.2.
Analytical methods and reagents
Samples were collected into 40 mL glass amber bottles containing 40 mg of sodium azide (preservative). Analysis of trace contaminants was determined based on a previously published rapid on-line solid phase extraction (SPE) and LC/MS/MS technique (Trenholm et al., 2009). Briefly, sample extraction and analysis of 1.5 mL was performed using a Symbiosis (Spark Holland) automated SPE, coupled to API4000 QTRAP (ABSCIEX) mass spectrometer. Oasis HLB cartridges were used for the online SPE. Separation was performed using a C18 column (Phenomenex) and with a mobile phase consisting of 5 mM ammonium acetate in DI water: methanol gradient. All samples were analyzed using positive electrospray ionization (ESI) and tandem mass spectrometry, or multiple reaction monitoring (MRM). Quantitation was performed using isotope dilution. For NDMA analysis, samples were collected into 1 L amber bottles, containing 1.0 g of sodium azide and 80 mg of sodium thiosulfate (for quenching any residual chlorine). NDMA quantitation was performed using a GC-MS/MS system and isotope dilution, using a method developed at the SNWA laboratories (Holady et al., 2012). NDMA standards were purchased from Ultra Scientific (Kingstown, RI, USA) and isotopically labeled NDMA was purchased from Cambridge Isotope Laboratories (Andover, MA, USA). For evaluating NDMA-formation potential,
samples were collected from the pilot system and spiked with preformed monochloramine before storing for 10 days at room temperature based on a previously published method (Mitch and Sedlak, 2004). Blank samples of deionized water were always below the method reporting limit (MRL) for NDMA and did not yield any measurable NDMA during the 10-day formation potential test. A monochloramine stock solution was prepared by rapidly mixing sodium hypochlorite into ammonium chloride solution following method described previously (Kumar and Margerum, 1987). Sodium hypochlorite, 10e14 wt% FAC, was obtained from VWR and standardized using iodometric titration prior to use. Ammonium chloride, 99% was obtained from Sigma Aldrich (St. Louis, MO, USA). In the NDMA destruction experiments with MBR filtrate, an NDMA spike solution was fed into the HiPOx with an in-line static mixer. As mentioned earlier, the NDMA was dissolved in water to prevent the introduction of any solvents that would contribute to OH-scavenging. For the CRW NDMA destruction experiment, NDMA was spiked at 2300 ng/L into a 3000 gallon tank. Ozone doses of 0.6, 1.5, 3, 6, and 10 mg/L were used with and without excess hydrogen peroxide at a 0.7 mol ratio to evaluate NDMA destruction. For the CRW TOrC destruction experiment, selected organic compounds were spiked into the batch tank at approximately 1000 ng/L, except for TCEP and TCPP which were spiked at approximately 5000 ng/L. For DOC analysis, samples were collected into glass vial and acidified to pH
